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Draft Report 9/24/07 for SAB C-VPESS Public Teleconferences on October 15 and 16,2007

This draft is a work in progress, does not reflect consensus advice or recommendations, has not been
reviewed or approved by the chartered SAB, and does not represent EPA policy.

Table of Contents for Draft SAB C-VPESS Report "Valuing the Protection of

Ecological Systems and Services"

1	INTRODUCTION	6

2	CONCEPTUAL FRAMEWORK	10

2.1.	An Overview of Key Concepts	10

2.1.1.	The Concept of Ecosystem Services	10

2.1.2.	The Concepts of Value	12

2.1.3.	The Concept of Valuation and Different Valuation Methods	18

2.2.	Ecological Valuation at EPA	23

2.2.1.	Policy Contexts at EPA Where Ecological Valuation Can be Important	23

2.2.2.	Institutional and Other Issues Affecting Valuation at EPA	25

2.2.3.	An Illustrative Example of Economic Benefit Assessment Related to Ecological
Protection at EPA	27

2.3.	An Integrated and Expanded Approach to Ecosystem Valuation: Key Features
32

2.3.1.	Early Consideration of Effects that are Socially Important	33

2.3.2.	Predicting Ecological Changes in Value-relevant Terms	34

2.3.3.	Drawing on Multiple Methods for Characterizing Values	36

2.4.	Steps in Implementing the Proposed Approach	38

2.5.	Conclusions and Recommendations	42

3	BUILDING A FOUNDATION FOR ECOLOGICAL VALUATION: PREDICTING
EFFECTS ON ECOLOGICAL SYSTEMS AND SERVICES	45

3.1.	The Road Map: A Conceptual Model	45

3.2.	Operationalizing the Conceptual Model: The Role of Ecological Production
Functions	48

3.3.	Challenges in Implementing Ecological Production Functions	51

3.3.1.	Understanding and Modeling the Underlying Ecology	51

3.3.2.	Identifying Ecosystem Services	55

3.3.3.	Mapping Changes in Ecological Inputs to Changes in Ecological Services	60

3.4.	Strategies to Provide the Ecological Science to Support Valuation	61

3.4.1.	Use of Indicators	61

3.4.2.	Use of Meta-analysis	65

3.4.3.	Opportunities regarding ecological data	66

3.5.	Directions for Ecological Research to Support Valuation	68

3.6.	Conclusions/Recommendations	68

4	METHODS FOR ASSESSING VALUE	71

4.1.	An Expanded Set of Methods	71

4.1.1.	Biophysical Ranking Methods	 76

4.1.2.	Ecosystem Benefit Indicators	 77

4.1.3.	Measures of Attitudes, Preferences, and Intentions	 78

4.1.4.	Economic Methods	 79

4.1.5.	Group Expression of Values and Social/Civic Valuation	81

4.1.6.	Decision Science Methods	82

4.1.7.	Methods Using Cost as a Proxy for Value	83

4.1.8.	Summary and Recommendations	84

4.2.	Value Transfer	90

4.2.1.	Transfer of Economic Benefits	90

4.2.2.	Transfer Methods	91

4.2.3.	Guidance Regarding Benefits Transfer	93

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Draft Report 9/24/07 for SAB C-VPESS Public Teleconferences on October 15 and 16,2007

This draft is a work in progress, does not reflect consensus advice or recommendations, has not been
reviewed or approved by the chartered SAB, and does not represent EPA policy.

4.2.4.	Screening Process	97

4.2.5.	Recommendations	98

5	CROSS-CUTTING ISSUES	99

5.1.	Analysis and Representation of Uncertainties in Ecological Valuation	99

5.1.1.	Introduction	99

5.1.2.	Sources of Uncertainty in Ecological Valuations	 100

5.1.3.	Approaches to assessing uncertainty	 101

5.1.4.	Using Uncertainty Assessment to Guide Research Initiatives	 104

5.2.	Communication of Ecological Valuation Information	105

5.2.1.	Applying General Communication Principles to Ecological Valuation	106

5.2.2.	Special Communication Challenges Related to Ecological Valuation	107

5.2.3.	Communicating Uncertainties and Ecological Valuation	 110

5.2.4.	Recommendations	112

6	APPLYING THE APPROACH IN THREE EPA DECISION CONTEXTS	114

6.1.	VALUATION FOR NATIONAL RULEMAKING	 114

6.1.1.	Introduction	 114

6.1.2.	Implementing the Proposed Approach	 116

6.1.3.	Conclusions	 132

6.2.	VALUATION FOR SITE-SPECIFIC DECISIONS	142

6.2.1.	Introduction	 142

6.2.2.	Opportunities for using valuation to inform remediation and redevelopment decision.
143

6.2.3.	Recommendations and discussion of valuation through illustrative site-specific
examples	 148

6.2.4.	Summary of recommendations for valuation for site-specific decisions	166

6.3.	VALUATION IN REGIONAL PARTNERSHIPS	 167

6.3.1.	EPA Role in Regional-scale Value Assessment	 167

6.3.2.	Case Study: Chicago Wilderness	 168

6.3.3.	Other Case Studies: Portland, Ore.; and the Southeast Region	 183

6.3.4.	Summary and Recommendations	186

1 SUMMARY OF MAJOR RECOMMENDATIONS AND CONCLUSIONS	190

APPENDIX A: SPECIAL TERMS AND THEIR USE IN THIS REPORT	198

APPENDIX B: DISCUSSION OF METHODS	199

BIOPHYSICAL RANKING METHODS	 199

Conservation Value Method	200

Rankings Based on Energy and Material Flows	208

ECOSYSTEM BENEFIT INDICATORS	215

MEASURES OF ATTITUDES, PREFERENCES, AND INTENTIONS	223

Brief description of the Methods	228

Relation of Methods to the C-VPESS Expanded and Integrated Assessment Framework	241

Status of Methods	244

Limitations	245

Treatment of Uncertainty	248

Research needs	249

ECONOMIC METHODS	253

Overview	253

Market-Based Methods	256

Non-market Methods - Revealed Preference	259

Travel cost	260

Hedonics	263

Averting behavior models	266

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This draft is a work in progress, does not reflect consensus advice or recommendations, has not been
reviewed or approved by the chartered SAB, and does not represent EPA policy.

Non-market Methods - Stated Preference	269

Combining Revealed and Stated Preference Methods	275

GROUP EXPRESSION OF VALUES AND SOCIAL/CIVIC VALUATION	277

Focus Groups	281

Referenda and Initiatives	284

Citizen Valuation Juries	296

DELIBERATIVE PROCESSES	305

Mediated Modeling	305

Valuation by Decision Aiding	314

METHODS USING COST AS A PROXY FOR VALUE	323

Replacement Costs	324

Tradable Permits	327

Habitat Equivalency Analysis	328

APPENDIX C: SURVEY ISSUES FOR ECOLOGICAL VALUATION: CURRENT BEST
PRACTICES AND RECOMMENDATIONS FOR RESEARCH	336

Defining Survey Research	336

Designs of Surveys	337

Elements of a Well-Defined Survey	339

Assessing Survey Accuracy	344

Challenges in Using Surveys For Ecosystem Protection Valuation	347

REFERENCES	353

ENDNOTES	385

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This draft is a work in progress, does not reflect consensus advice or recommendations, has not been
reviewed or approved by the chartered SAB, and does not represent EPA policy.

Lists of Figures, Tables, and Text Boxes
List of Figures

Figure 1: Components of Ecological Valuation	32

Figure 2: Process for Implementing an Expanded and Integrated Approach to Ecological

Valuation	40

Figure 3: Illustration from Covich et al., 2004, Showing Relationships of Major

Functional Types to Ecological Services	46

Figure 4: Graphical Depiction of Ecological Production Functions	50

Figure 5: Indicators of Ecological Properties at Different Levels of Organization	63

Figure 6: General Overview of the Impact of CAFOs	118

Figure 7: Framework for Net Environmental Benefit Analysis (from Efoymson et al.,

2003)	 145

Figure 8: Integration of Valuation information with traditional process to achieve

improved performance	146

Figure 9: Visualization of Forest Conditions and Actual Photos from Ribe et al. (2002)

	233

Figure 10: Graphical Representation of Ecosystem Service Loss and Recovery through
Natural and Active Restoration Over Time	328

Tables

Table I. A Classification of Concepts of Value as Applied lo l-colouical Systems and

Their Services	16

Table 2: Introduction to Methods Assessed by the Committee	72

Table 3: Table Summarizing Methods Discussed in this Report	86

Table 4: Table of Alternative Unit Value Transfers	95

Table 5: Table of Qualitative Discussions of Potential Ecological Effects of Atmospheric

Pollutants Discussed in the First Prospective Benefit Cost Analysis (1999)	 139

Table 6: Ecosystem Service Matrix for Leviathon Mine (from Wilson, 2004)	 161

Table 7: Example Items from Survey Supporting USDA Forest Service Strategic Plan

for 2000 required by the Government Performance and Results Act	230

Table 8: Facsimile of Illustrative Choice Questions from Chattopadhyay et al. (2005)236
Table 9: Comparative Matrices of Attributes for Three Hypothetical Decision-Aiding
Valuation Scenarios	317

List of Text Boxes

Text Box 1
Text Box 2
Text Box 3
Text Box 4
Text Box 5
Text Box 6
Text Box 7

The Challenge of Choosing a Unit Value for Economic Benefits Transfer 94

The Aquaculture Effluent Guidelines	135

The CAFO Effluent Guidelines	137

The Prospective Economic Benefits of the Clean Air Act Amendments .138

Net Environmental Benefit Analysis	144

Charles George Landfill	148

DuPage County Landfill	149

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Text Box 8: Avtex Fibers Site	150

Text Box 9: Leviathan Mine Superfund Site	150

Text Box 10: Possible Ecological Impacts and Provision of Services from the Protection

or Restoration of Watersheds Based on the Work of Chicago Wilderness	174

Text Box 11: National Telephone Survey	229

Text Box 12: Perceptual Surveys	232

Text Box 13: Conjoint Surveys	235

Text Box 14: Direct Analysis of Public Decisions to Accept Pollution or Resource

Depletion	285

Text Box 15: Referendum/Initiative Analysis Followed by a Survey	285

Text Box 16: Public Decisions to Accept Pollution or Resource Depletion Followed by a

Survey	286

Text Box 17 Referenda and Initiatives Used to Validate Contingent Valuation	292

Text Box 18: A Valuation Exercise Illustrating Use of Citizen Juries	299

Text Box 19: Types of Attributes	316

Text Box 20: Equation for Habitat Equivalency Analysis	329

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This draft is a work in progress, does not reflect consensus advice or recommendations, has not been
reviewed or approved by the chartered SAB, and does not represent EPA policy.

1 INTRODUCTION

The Environmental Protection Agency's (EPA's) mission is to protect human health and the
environment. During its history EPA has focused much of its decision-making expertise on the first
part of this mission, in particular the risks to human health from chemical stressors in the
environment. Although protecting human health is the bedrock of EPA's traditional expertise, the
broad mission of the EPA goes beyond this. In fact, EPA's Strategic Plan (U.S Environmental
Protection Agency 2006b) explicitly identifies the need to ensure "healthy communities and
ecosystems" as one of its five major goals. In addition, EPA's efforts in protecting ecological
resources—and its authority for doing so—have been documented in Agency publications and
independent historical sources (U.S. Environmental Protection Agency 1994;U.S. Environmental
Protection Agency Risk Assessment Forum 2003; U.S. Environmental Protection Agency Science
Advisory Board 2000; Hays 1989; Russell III 1993).

EPA's mission to protect the environment requires that the Agency understand and protect
ecological systems. "Ecosystem" is the term used by ecologists to describe the dynamic complex of
plant, animal, and microorganism communities and the non-living environment interacting as a
system. For example, a forest ecosystem is comprised of the trees in the forest plus the birds, insects,
soil microorganisms, and streams that inhabit or run through it. Ecosystems provide basic life
support for human and animal populations and are the source of spiritual, aesthetic, and other human
experiences that are valued in many ways by many people. There has been a growing recognition of
the numerous and varied services that ecosystems provide to human populations through a wide
range of ecological functions and processes (see, for example, Daily 1997). Ecosystems provide not
only goods and services directly consumed by society such as food, timber, and water, but also
services such as flood protection, disease regulation, pollination, and disease, pest, and climate
control. In addition, there is increasing recognition of the impact of human activities on ecosystems
(see, for example, Millennium Assessment). Examples of this impact include not only traditional air
and water pollution (such as sulfur dioxide emissions, ground-level ozone, and eutrophication), but
also land conversions that lead to deforestation or loss of wetlands and biodiversity; global warming;
changes in the nitrogen cycle; invasive species; and aquifer depletion.

Given the vital role that ecosystems play in our lives, changes in the state of these systems or
the flow of services they provide can have important implications. EPA actions (e.g., regulations,

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rules, programs, policy decisions) can be one source of these changes. Many EPA actions relating to
air quality, water quality, and land use affect the condition of the environment and the flow of
ecological services from it. These impacts can occur narrowly at a local scale or broadly at a
national scale.

Despite their importance, the ecological impacts of EPA actions have received relatively
limited consideration in EPA policy analyses. Rather, these analyses have tended to focus on a
relatively narrow set of ecological endpoints, such as those identified by tests required for pesticide
regulation (e.g., the effects on survival, growth, and reproduction of aquatic invertebrates, fish, birds,
mammals, and both terrestrial and aquatic plants) or mortality to fish, birds, plants, and, animals, as
required by provisions of several laws administered by the Agency1 (U.S. Environmental Protection
Agency Risk Assessment Forum 2003). Given EPA's responsibility to ensure healthy communities
and ecosystems the Agency must consider the full range of impacts that its actions will have not only
on human health but also on individual organisms and plant and animal populations, as well as on
the key structural and functional characteristics of communities and ecosystems.

To promote good decision making, policy makers also require information about how much
ecosystems contribute to society's well-being. This need is increasingly recognized both within and
outside the Agency. The stated goal of EPA's recently released Ecological Benefits Assessment
Strategic Plan (EBASP) is to "help improve Agency decisionmaking (sic) by enhancing EPA's
ability to identify, quantify, and value the ecological benefits of existing and proposed policies" (p.
xv). In addition, information about the value of ecosystems and the associated impacts of EPA
actions can help inform the public about the need for ecosystem protection and the extent to which
specific policy alternatives address that need.

Despite EPA's stated mission and mandates, a gap exists between the need for understanding
and protecting ecological systems and services and EPA's ability to address this need. This report is
a step toward filling that gap. It describes how an integrated and expanded approach to ecological
valuation might help the Agency describe and measure the value of protecting ecological systems
and services, thus better meeting its overall mission.

This report was prepared by the Committee on Valuing the Protection of Ecological Systems

and Services (C-VPESS), which was formed by EPA's Science Advisory Board (SAB) in 2003. The

SAB saw a need to complement the Agency's ongoing work by offering advice on how EPA might

better value the protection of ecological systems and services and how that information could

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Draft Report 9/24/07 for SAB C-VPESS Public Teleconferences on October 15 and 16,2007

This draft is a work in progress, does not reflect consensus advice or recommendations, has not been
reviewed or approved by the chartered SAB, and does not represent EPA policy.

support decision making to protect ecological resources. Toward this end, it formed C-VPESS,2 a
group of experts from the disciplines of decision science, ecology, economics, engineering,
philosophy, and psychology, with an emphasis on ecosystem protection. The committee's charge
was to undertake a project designed to improve the Agency's ability to value ecological systems and
services.3 The SAB set the following goals for this project: a) assessing Agency needs for valuation
to support decision making; b) assessing the state of the art and science of valuing protection of
ecological systems and services, and c) identifying key areas for improving knowledge,
methodologies, practice, and research at EPA.

This report provides advice for strengthening the Agency's approaches for valuing the
protection of ecological systems and services, facilitating the use of these approaches by decision
makers, and identifying the key research areas needed to bolster the science underlying ecological
valuation.4 It identifies the need for an expanded and integrated approach for valuing EPA's efforts
to protect ecological systems and services. It provides advice to the Administrator, EPA managers,
EPA scientists and analysts, and EPA staff across the Agency concerned with ecological protection.
It adopts a broad view of EPA's work, which it understands to encompass national rule making,
regional decision making, and programs in general that protect ecological systems and services. It
recommends that EPA expand its current approach in important ways.

This report appears at a time of lively interest internationally, nationally, and within EPA in
the issue of valuing the protection of ecological systems and services. Since the establishment of the
SAB C-VPESS, a number of major reports have focused on ways to improve the characterization of
the important role of ecological resources (Millennium Ecosystem Assessment Board 2003; Silva
and Pagiola 2003; National Research Council 2004; Pagiola, von Ritter et al. 2004; Millennium
Ecosystem Assessment 2005). In addition, the Agency itself has engaged in efforts to improve
ecological valuation. The most recent product of these efforts is the EBASP report noted above
(U.S. Environmental Protection Agency 2006a). This report examines EPA efforts to improve
ecological valuation, which have been geared toward the use of economic valuation in benefit-cost
analysis. EPA also has sought to strengthen the science supporting ecological valuation through the
extramural Science to Achieve Results (STAR) grants program. STAR grants involving ecological
valuation have primarily applied economic valuation methods to various ecosystem services.

The committee has both learned from and built upon these recent efforts. The C-VPESS
distinguishes its work from the earlier efforts, however, in several key ways. First, the C-VPESS

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This draft is a work in progress, does not reflect consensus advice or recommendations, has not been
reviewed or approved by the chartered SAB, and does not represent EPA policy.

considers EPA the principal audience for its work. In particular, it analyzes ways in which EPA can
value its own contributions to the protection of ecological systems and services, so that the Agency
can make better decisions in its eco-protection programs. Many of the recent studies (for example,
the Millennium Assessment and NRC report) do not consider the specific policy contexts or
constraints faced by EPA. Second, most previous work has concentrated on economic valuation as
the primary valuation method. C-VPESS, by contrast, is inter-disciplinary and does not focus solely
on economic methods or values. The committee will offer advice on several approaches to
characterizing or estimating values and in each case will emphasize issues relevant to EPA policy
and decision-making.

The report is structured as follows. Chapter 2 provides an overview of the conceptual
framework and general approach advocated by the committee. It discusses fundamental concepts as
well as the current state of ecological valuation at EPA. Most importantly, it identifies the need for
an expanded and integrated approach to ecological valuation at EPA and describes the key features
of this approach. Subsequent chapters develop the basic principles outlined in Chapter 2 in more
detail, with a focus on implementation. Chapter 3 discusses the part of the implementation process
related to prediction of changes in ecological systems and services that stem from EPA actions or
decisions. Chapter 4 then discusses a variety of methods for valuing these changes. Cross-cutting
issues relating to uncertainty and communication are discussed in Chapter 5 Recognizing that
implementation of the process may vary depending on the decision context, Chapter 6of the report
discusses implementation in three specific contexts where ecological valuation could play an
important role in EPA analysis: national rulemaking, site-specific decisions (regarding, for example,
cleanup and restoration), and regional partnerships. Finally, Chapter 7provides a summary of the
report's major conclusions and recommendations..

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2 CONCEPTUAL FRAMEWORK
2.1. An Overview of Key Concepts

2.1.1. The Concept of Ecosystem Services

As noted above, the term ecosystem describes a dynamic complex of plant, animal, and

microorganism communities and the non-living environment, interacting as a system.

Ecosystems encompass all organisms within a prescribed area including humans, who are often
the dominant organism. Ecosystem functions or processes are the characteristic physical,
chemical, and biological activities that influence the flows, storage, and transformation of
materials and energy within and through ecosystems. These activities include processes that link
organisms with their physical environment (e.g., primary productivity and the cycling of
nutrients and water) and processes that link organisms with each other, indirectly influencing
flows of energy, water and nutrients (e.g., pollination, predation and parasitism). These
processes in total describe the functioning of ecosystems.

Ecosystem services are the direct or indirect contributions that ecosystems make to the
well-being of human populations. Ecosystem processes and functions contribute to the provision
of ecosystem services; however, they are not synonymous with ecosystem services. Ecosystem
processes and functions describe biophysical relationships and exist whether or not humans
benefit from them. These relationships only generate ecosystem services, though, if they
contribute to human well-being. Thus, ecosystem services cannot be defined independently of
human values.

The following categorization of ecosystem services has been used by the Millennium
Ecosystem Assessment:

a)	Provisioning services - services from products obtained from ecosystems. These
products include food, fuel, fiber, biochemicals, genetic resources, and fresh
water. Many, but not all, of these services are traded in markets.

b)	Regulating services - services received from regulation of ecosystem processes.
This category includes services that improve human well-being by regulating the
environment in which people live. These services include flood protection, human
disease regulation, water purification, air quality maintenance, pollination, pest

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control, and climate control. These services are generally not marketed but many
have clear value to society.

c)	Cultural services - services that contribute to the cultural, spiritual, and aesthetic
dimensions of people's well-being. They also contribute to establishing a sense
of place.

d)	Supporting services - services that maintain basic ecosystem processes and
functions such as soil formation, primary productivity, biogeochemistry, and
provisioning of habitat. These services affect human well-being indirectly by
maintaining processes necessary for provisioning, regulating, and cultural
services.

This categorization suggests a very broad definition of services, limited only by the
requirement of a contribution (direct or indirect) to human well-being. This broad approach
reflects the recognition of the myriad ways in which ecosystems support human life and
contribute to human well-being. Alternatively, Boyd and Banzhaf (2006) propose a definition
that focuses on services as "end products of nature", i.e., "components of nature, directly
enjoyed, consumed or used to yield human well-being" [emphasis added]. They stress the need
to distinguish between intermediate products and final (or end) products and include only final
outputs in the definition of services, since these are what affect people most directly and are
consequently what they are most likely to understand. Under this definition, ecosystem
functions and processes, such as nutrient recycling, are not considered services; while they
contribute to the production of ecological end products or outputs, they are not outputs
themselves.5 Principles for defining ecosystem services are discussed in more detail in Chapter 3
of this report.

Regardless of the specific definition used, the general concept of ecosystem services
plays a key role in evaluating policies related to ecological protection. Even without any
subsequent valuation, explicitly listing the services derived from an ecosystem - and using the
best available methods in the ecological, social, and behavioral sciences to develop that list - can
help to ensure appropriate recognition of the full range of potential impacts of a given policy
option. This can help make the analysis of ecological systems more transparent and accessible
and can help inform decision makers of the full range of potential impacts stemming from
different options before them.

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The concept of ecosystem services provides an approach to evaluating the many ways in
which ecological systems, and changes to those systems induced by human actions, affect human
well-being. Ecosystems, however, can also be valued for reasons that are independent of effects
on human well-being. As discussed in the following section, the committee recognizes that
ecosystems can be important not only because of the services they provide to humans directly or
indirectly, but also for other reasons including respect for nature based on moral, religious, or
spiritual beliefs and commitments.

2.1.2. The Concepts of Value

The committee recognizes that there are many sources or types of value that are relevant

when valuing the protection of ecosystems and their services. In considering concepts of value,

a fundamental distinction can be made between those things that we value as ends or goals and

those things that we value only as means. To value something as a means is to value it for its

usefulness in helping to bring about an end or goal that is valued in its own right. Things or

actions valued for their usefulness as means in this sense are said to have instrumental value.

Alternatively, something can be valued for its own sake as an independent end or goal. While a

possible goal is "maximizing human well-being," one could envision a range of other possible

social goals or ends including "protecting biodiversity," "sustainability," or "protecting the

health of children." Things valued as ends are sometimes said to have "intrinsic value." This

term has been used extensively in the philosophical literature but there is not general agreement

on its exact definition.6

The distinction between ends and means plays an important role when thinking about

valuing ecological systems and services. People have material, moral, religious, aesthetic, and

other interests, all of which can affect their thoughts, attitudes, and actions toward nature in

general and, more specifically, toward ecosystems and the services they provide. Thus, when

people talk about environmental values, the value of nature, or the values of ecological systems

and services, they may have different things in mind (e.g., ends vs. means). For example, some

people claim that the very existence of a species or ecological system has value in itself in

addition to any instrumental value derived from the usefulness of the services it provides. This

claim can be interpreted in different ways. It could mean that the existence of a species or an

ecological system is valuable because people derive satisfaction from its existence, independent

of specific uses they may make of its services. Economists would interpret this type of value as

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"existence value", which is a form of instrumental value since it is based on the premise that the
existence of the species or ecological system is one of many things that contribute to human
well-being. Alternatively, the claim could be interpreted to mean that an ecological system is
valuable as an end or goal for its own sake, implying that the reasons for this claim are
independent of the contribution that the existence of the ecological system can make to human
well-being. This interpretation of the claim is consistent with values in which the existence or
well-being of other species or the state of ecosystems can be ends in themselves.

The committee recognizes that ecosystems can be valued both as independent ends or
goals and as instrumental means to other ends or goals. To reflect this recognition, throughout
this report, the term value is broadly used. It includes values that stem from contributions to
human well-being as well as values that reflect other considerations, such as social and civil
norms (including rights), and moral, religious, and spiritual beliefs and commitments.

Recognizing that values can be instrumental or intrinsic, this report next turns to how
those values can be defined. A key implication of instrumental values is substitutability.
Substitutability means that more of one thing can be traded off against less of something else as
long as both contribute to achieving the same goal. Assuming there is more than one thing that
contributes to the achievement of a goal and that alternative means are substitutable, the
instrumental value of something can be defined as the amount of something else that would
make an equivalent contribution to the goal and could replace the thing in question if it were to
be lost. For an example taken from economic valuation methods, if the end goal is the
maximization of human well-being and both the existence of a species and money contribute to
that goal, then the value of the species can be defined as the amount of money that would be
needed to offset the loss in human well-being that would result from loss of the species.
Likewise, if the end goal is the provision of clean water to a given community and this goal can
be achieved through either watershed protection or the construction of a water purification plant,
then given that the clean water from either source is accepted as equivalent the value of
watershed protection can be defined as the cost savings from not having to build the purification
plant.

While the definition of instrumental value is clear, it is less clear how to define, measure,
and ultimately quantify intrinsic value. When something is an end in itself, its value cannot be

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determined in terms of trade-offs since there are no substitutes for something that is an end in
itself. For example, if ecosystems are viewed as ends in themselves or are valued for other than
human utilitarian purposes (e.g., out of respect for, or acceptance of, ethical obligations toward
nature), then a water purification plant cannot be substituted for watershed protection. Defining
the value of items as their contribution toward achieving a goal (Costanza 2000) requires that the
identification of the goal be separate from the item or good being valued (i.e., separate from the
means for attaining the goal). With intrinsic values, this is not possible since the good being
valued and the goal are not separate. In this sense, intrinsic values cannot be quantified or
measured. Nonetheless, as envisioned by the committee in this report, identifying and providing
information about intrinsic goals relating to ecosystem protection, including measures of how
strongly people care about them (perhaps relative to other goals), is an important component of
the assessment of ecosystem values and a legitimate consideration for Agency decision making.

This raises the question of how these intrinsic values can be compared to instrumental
values when tradeoffs are, in fact, required. In other words, how does society balance these
intrinsic values - moral, aesthetic, religious or spiritual goals - with its interests in instrumental
contributions to human well-being, both as individuals and as a society? This cannot be done by
a direct comparison of the associated values. Rather, if trade-offs are required, society must
engage in political and deliberative discussion of alternative goals and visions for the future in
order to balance intrinsic and instrumental values. This discussion should be an ongoing and
vital part of any democratic society.

Although the concepts of instrumental and intrinsic value provide a broad categorization
of values, other distinctions between different types or concepts of value can also be made and
are important for understanding the values associated with ecological systems and services. For
example, values can be classified as either anthropocentric values or nonanthropocentric values.
Anthropocentric values are based on the contributions that ecological systems and services make
to human well-being. Nonanthropocentric values are based on a variety of ethical and
philosophical perspectives. This category includes biocentric and eco-centric values, which are
based on an evaluation of ecological changes and their effects on ecosystems or nonhuman
species, and values stemming from theories of value that are not based directly on human well-
being. Note that the anthropocentric values derived from contributions of ecosystem services to
human well-being are often referred to as the "benefits" from ecosystem services (see

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Millennium Assessment). The term benefits, however, has a very precise meaning in the context
of EPA regulatory impact analyses conducted under OMB guidance (see further discussion that
follows). In that context, benefits are defined by the economic concept of willingness to pay for
a good or service or willingness to accept compensation for it. Thus, the term "benefits" means
different things in different contexts. For this reason, throughout this report the committee refers
to the broader concept of anthropocentric values as contributions to human welfare, and uses the
term "benefits" only when there is no potential for confusion about what it includes.

In addition to the distinction between anthropocentric and non-anthropocentric values,
values can also be distinguished by whether they are preference-based or bio-physical. Values
based directly on human preferences can be either instrumental or intrinsic values and can be
either anthropocentric or non-anthropocentric. In contrast, bio-physical values do not directly
reflect human preferences. However, they can still be either implicitly anthropocentric or non-
anthropocentric. They are non-anthropocentric when they reflect intrinsic values unrelated to
human well-being; and they are implicitly anthropocentric when they reflect a prior decision or
commitment to a bio-physical goal that is deemed to be important for human welfare. For
example, values based on contributions to a goal of preserving biodiversity can reflect either a
belief that biodiversity preservation has intrinsic value (a non-anthropocentric approach), or a
prior commitment to preserving biodiversity because of its importance for human welfare (an
implicitly anthropocentric approach). In either case, the value of an ecosystem change is
defined in terms of its contribution to the goal of preserving bio-diversity, which does not require
direct information about people's preferences for that particular change. Similarly, if society has
identified a goal of ensuring clean water to a community (an anthropocentric goal), then the
contribution of watershed protection to that goal can be valued without direct information about
human preferences.

The discussion above highlights the fact that there are many concepts of value and
alternative ways to categorize them. Table 1 lists the various concepts of value that the
committee has considered in its deliberations, categorized as preference-based versus bio-
physical. While this is not the only way to categorize values, it is one that has proven useful for
the committee. What follows is a brief description of the major features of each of these
concepts of value. Note that these value concepts are not mutually exclusive. For example,
values expressing attitudes or judgments can be based on the same self-interested utilitarian

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goals as those underlying the concept of economic values, or on the considerations that underlie
social/civic values. Likewise, preferences that are constructed can relate to self-interested
attitudes or judgments as well as expressed social/civic values.

T.ihle I: A Cliissiriciilion of Cnneepls of \';iliie ;is Applied lo l-'.eolu^ieiil S\stems ;iihI Their Sen iees

Preference-based Values'

A Ixonomic Values
1} Constructed Preferences
(' Community-leased or Social Ci\ic Values
I). Attitudes or Judgments

ISio-nlnsical Values

A liio-ecolouical Values
1} I Jieiu\-haseel Values

Economic values are based on individuals' preferences and assume that individuals are
self-interested and that they should be allowed to value goods and services based on their
judgment of the contribution those goods and services make to their own well-being or utility
(the concept of consumer sovereignty, Freeman, 2003). People are assumed to be rational actors
and have well-defined and stable preferences over alternative outcomes. In addition, the choice
of one outcome over another is assumed to imply that the chosen outcome was judged to result in
a higher level of well-being for the individual, consistent with the principle of consumer
sovereignty. Economic values can include both use and nonuse values. They are based on a
coherent theory of welfare economics and identify the tradeoffs that individuals are willing to
make, given their income and the prices they face. They are normally expressed in monetary
units and allow a comparison of the values of ecosystem services with the values of other
services produced through environmental policy changes (for example, effects on human health)
and with the costs of those policies.

In contrast to the assumption underlying economic values, some researchers have argued
that, particularly when confronted with unfamiliar choice problems, individuals do not have

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well-formed preferences and hence values, implying that simple statements of preferences or
willingness to pay are unreliable (Gregory and Slovic, 1997; Lichtenstein and Slovic, 2006).
These authors have advocated using a structured or deliberative process as a way of assisting
respondents in learning about the ecological services to be valued and in constructing their
preferences and values. This report refers to values arrived at by these processes as constructed
values. The difference between economic values and constructed values can be likened to the
difference between the work of an archeologist and that of an architect. Economic methods
assume preferences exist and simply need to be "discovered" (implying the analyst works as a
type of archeologist), while deliberative methods assume that preferences need to be built
through the valuation process (similar to the work of an architect). As a result, the values
expressed by individuals (or groups) engaged in this process are expected to be influenced by the
process itself. Constructed values can include both individual values (reflecting self-interest)
and community or social/civic values.

Community-based or social/civic values are based on the assumption that, when placed in
a position of making choices about public goods (goods that when made available to one person
are available to all), individuals make their choices based on what they think is good for society
as a whole rather than what is good for them as individuals. In other words, people base their
choices on their conception of social preferences or community-based preferences rather than
their own self-interest. In this case, individuals could place a positive value on a change that
would reduce their own individual well-being (see, for example, Jacobs 1997, Costanza and
Folke 1997, or Sagoff 1998).

Attitude or judgment-based values are based on empirically derived descriptive theories
of human attitudes, preferences, and behavior. In contrast to economic values, preferences are
not expressed in terms of willingness-to-pay (or accept) and they are not typically constrained by
income or prices, especially those that are outside the context of the specified assessment
process. Rather, the values are derived from individuals' judgments of relative importance,
acceptability, or preferences across the array of changes in ecosystems or services presented in
the assessment. Preferences and judgments are often expressed through responses to surveys
(e.g., choices, ratings or other indicators of importance). The basis for judgments may be
individual self-interests, community well-being, or accepted civic, ethical, or moral obligations
relevant to ecosystems and ecosystem services. Moreover, emotions and intuitions are accepted

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as having equal and often greater influence on value-relevant judgments and preferences than
rational processes.

All of the above concepts of value are based directly on human preferences.
Alternatively, bio-ecological values are defined in a way that does not depend directly on human
preferences; rather, they reflect the contribution of a change to a pre-specified ecological or
conservation goal (e.g., species or biodiversity preservation). As noted above, this goal can
reflect intrinsic values (e.g., a biocentric or ecocentric view) or an underlying assumption or
prior decision based on instrumental value (i.e., a belief that biodiversity is important for human
well-being). Bioecological values are based on known or assumed relationships between
targeted ecosystem conditions (e.g., biodiversity, biomass, energy transfer, and transformation)
and ecosystem functions. For example, the value of changes in biodiversity could be defined in
several different ways, including individual measures such as genetic distance or species
richness, as well as more comprehensive measures that reflect multiple ecological
considerations. What levels of bioecological measures are deemed better or worse in a given
policy context may be determined solely on biological grounds (a biocentric approach) or on the
basis of determined (or presumed) relationships to things people value.

Energy-based values, which reflect an energy theory of value, are based on the impact of
an ecological change on energy or materials flows into and out of ecological systems. They are
defined as the free energy (or "exergy") required directly and indirectly to produce a good or
service. While these values reflect human preferences indirectly, they were designed to provide
an alternative way to define value independently of human preferences.

As noted above, the committee considered all of these various types of value in its
deliberations. The committee's recommendations throughout this report reflect a recognition
that not only different individuals, but also different disciplines (e.g., decision science, ecology,
economics, philosophy, psychology), think of the concept of value in different ways. The
committee believes that recognizing this is an important first step in valuing the protection of
ecological systems and services.

2.1.3. The Concept of Valuation and Different Valuation Methods

The term "valuation" generally refers to the process of measuring either the value, or

change in value, of an ecosystem, its components, or the services it provides. The committee
believes that valuation should seek to characterize or measure the values actually generated by

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ecological systems, regardless of how well those values are currently perceived by the general
population. This is a broader conception of valuation than one often used in practice, where
assessments tend to focus on values currently perceived, and expressed by individuals in the
general population. As discussed below, in some cases, an ecological change may have
important implications that are not widely recognized or understood by the general public. For
example, Weslawski, et al. (2004) indicated that the invertebrate fauna found in soils and
sediments are important in remineralization, waste treatment, biological control, gas and climate
regulation, and erosion and sedimentation control. Yet, their analysis showed that the general
public had no understanding or appreciation of these services. They do have an appreciation of
the higher level services or end-point services, such a clean water and aesthetics, and, of course,
foods that could be derived from the system.

Regardless of the level of public understanding, valuation should seek to measure the
value of the actual impact rather than simply the perceived impact. Thus, valuation can be
viewed as providing a comparison of the predicted outcomes, based on the best available science,
under two alternative scenarios: having a specific, proposed policy in place or maintaining the
baseline or status quo. In valuing a change in ecosystem services, both the baseline before the
policy change and the alternative world with the policy change must be specified. Similarly,
when measuring the value of an ecosystem itself (rather than a change in that ecosystem), the
baseline is the world without that ecosystem, a world which might be difficult to describe in any
meaningful way. It is important to note, however, that although valuation should be informed by
the best available science, it ultimately seeks to reflect the values that would be held by a fully-
informed general public, not merely the personal values or preferences of scientists. Basing
valuation on the personal preferences of scientists rather than the general public would
undermine the usual presumption that, in a democratic society, the values held individually and
collectively within that society should be considered in public policy decisions, and that public
involvement is central to democratic governance (e.g., Berelson, 1952).

Just as there are many types of values, there are a number of valuation methods that can
be used for estimating or measuring values from ecosystems or services. Some of these methods
are well developed while others are in need of further development and testing in the context of
valuing the protection of ecosystems and services. Specific methods are discussed in detail in
Chapter 4 and Appendix B of this report. A key tenet of valuation as defined in this report is the

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need to explicitly identify the type(s) of value to be measured and the appropriate method(s) for
measuring those values. Methods differ on a number of dimensions, including the type(s) of
value they attempt to measure (and hence their theoretical foundations and assumptions), the
type(s) of metrics or outputs produced and whether they produce single or multiple metrics.

These differences need to be explicitly recognized and considered as part of the valuation
process.

As noted, different valuation methods express values in different ways, including
monetary units, biophysical units, or indices. Economic valuation methods typically use metrics
expressed in monetary units. Other social scientists and ecologists have developed measures or
indices expressed in a variety of non-monetary units such as relative preference or importance
ratings by samples of the general public or stakeholders, or biophysical indices calculated
through expert analyses. When these measures or indices are used to make judgments about
which outcomes are preferred, these measures are considered a form of non-monetary valuation.
For example, bioecological valuation methods might be used to value alternative landscape
management plans in terms of how well they do in conserving biodiversity, where landscape
management alternatives that conserve more biodiversity are considered to be more valuable.

When multiple methods are used to capture different sources of value, the question of
aggregation across methods arises. It is clear that values cannot be aggregated across methods
that yield value estimates in different units. However, even when units are comparable (e.g.,
both methods yield monetary estimates of value), aggregation across methods may not be
appropriate. Because of their different assumptions, the different methods can measure quite
different things and yield values that are conceptually different and hence not comparable. As a
result, simple aggregation across methods is generally not scientifically justified. For example, it
would be conceptually inconsistent to add monetary value estimates obtained from an economic
method and monetary estimates obtained from a citizen jury (or, alternatively, a deliberative
process in which preferences are constructed) since the two are not based on the same underlying
premises. Nonetheless, information about both estimates of value may be of interest to policy
makers and play a key role in policy decisions. In such cases, EPA should report value estimates
separately rather than seeking to aggregate across methods.

Aggregation issues also arise when considering alternative valuation methods. Some
methods involve aggregation across components of value and yield a single metric of the value

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of a particular ecosystem or ecological change, while others yield multiple metrics of value.
Valuation methods that seek to aggregate all components of value into a single metric, such as a
formal economic willingness-to-pay or willingness-to-accept analysis, must weight various
sources of value as part of the valuation process and report estimated aggregate values that
reflect these weights. In contrast, other methods do not seek to aggregate sources of value as part
of the assessment. Rather, they report the information about the various value dimensions
separately and allow decision makers to weigh these components more or less formally in the
process of coming to a decision. Methods that produce a single metric are not necessarily
preferred to those that do not. Which approach is more appropriate or useful will in general
depend on the decision context. For example, if the context requires a ranking or choice based
on a single criterion (e.g., net economic analysis of benefits and costs), then a valuation approach
that yields a single metric will be needed. In contrast, in a decision context where multiple
values are involved (e.g., human health, threatened species, aesthetics, social equity, and other
civil obligations) and decision makers themselves are charged with appropriately weighing and
balancing competing interests and resolving trade-offs, a multi-attribute approach will be
preferred. Depending upon the context, this weighing and balancing might be done through
political discourse or through a deliberative, decision-aiding process (see, for example, Clemen
1996; Arvai, Gregory, and McDaniels 2001; Arvai and Gregory 2003). It is important to note,
however, that in either case a decision ultimately requires, explicitly or implicitly, weighing and
making trade-offs among the multiple values. What varies among valuation approaches is where
in the decision making process the weighing of values is done and by whom.

Finally, some valuation methods, such as economic methods and socio-psychological
methods based on surveys, assume that (1) individuals know and can consistently express their
preferences, and (2) individuals are well informed about alternatives, at least those they face in
the assessment, and are aware of the potential consequences of the choices they make. These
assumptions can be problematic when it comes to applying these valuation methods to
ecosystems or services. First, for complex issues such as ecosystem protection, individuals are
not likely to be aware of or fully appreciate all of the ecosystem's contributions. For example,
although individuals might understand the recreational contributions to human well-being
associated with a given EPA action to limit nutrient pollution in streams and lakes, they might
not recognize or fully appreciate the associated nutrient cycling or water quality implications.

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As a result, the policy preferences or values they express through survey methods or through
their behavior will reflect that incomplete information. For example, individuals might respond
to a survey or behave as if they place no value on an ecosystem service if they are ignorant of the
role of that service in contributing to their well-being or other goals.

Second, as noted above, when people have limited information about and little experience
with an ecosystem or service, their preferences may not be well-formed and may be subject to
intentional or unintentional manipulation or bias through (e.g., as by changes in wording or
framing in surveys or by labeling or placement of items in retail stores. The extent to which this
is true is the subject of debate, and most likely varies with the context. (See a more detailed
discussion in Appendix B.) If preferences and values regarding ecological systems and services
are not well-formed, then they cannot be accurately measured or characterized by valuation
methods that assume well-formed preferences. For example, individuals can have strongly held
values that they find difficult, impossible or inappropriate to express in terms of monetary units.
If this is so, requiring individuals to express such values in monetary equivalents (as is typical in
economic valuation) may compel them to assume an individual consumer perspective that is
unfamiliar or even offensive in that context. When preferences are not well-formed, survey-
based methods, whether using willingness-to-pay or attitude ratings, may force the respondent to
construct their preferences from more basic values in the context of the valuation itself,
jeopardizing the validity of the values derived from those responses. Alternatively, and in many
cases preferably, the construction of people's preferences can be made explicit and facilitated
through use of a valuation method based on discourse and deliberation.

When considering alternative methods, policy makers should look for which of these
methods, or what combinations, might give the best assessment of the values of ecosystems and
services in particular policy contexts. In circumstances in which the individuals involved can be
expected to be well informed and to have well-formed preferences for the policy options and
outcomes in question, decision makers should put more weight on the stated and revealed
preferences of stakeholders and the public as measured by appropriate economic and social-
psychological methods. In circumstances in which individuals are likely to be ill-informed or to
have ill-formed preferences, policy makers should seek to ensure that individuals expressing
values have a sufficient understanding of the likely biophysical impacts of alternative policy
options and their implications for ecosystems and the services they provide. For example, in

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specific policy contexts, using deliberative methods such as mediated modeling (see Appendix
B) as part of the valuation process can help stakeholders better understand the ecological effects
of alternative choices. More generally, public agencies have an obligation to aggressively pursue
public education and involvement when a gap exists between public knowledge (and hence
expressed preferences) and scientific understanding.

2.2. Ecological Valuation at EPA

As noted in the introduction, in contrast to previous studies, this report is focused

specifically on ecological valuation within EPA. This necessitates consideration of some issues
that might not be considered in more general discussions of ecological valuation. The committee
recognizes that EPA operates in a variety of different decision contexts where valuation might be
useful. While much of the interest in ecological valuation at EPA has focused on valuation
needs in the context of national rule making, valuation can also be useful in other decision
contexts as well. The need for valuation arises in different parts of the Agency for different
purposes and for different audiences. Some of the needs present structured requirements for
valuing protection of ecological systems and services, while needs in other contexts are less
prescriptive. In addition, EPA faces institutional constraints that both influence and limit how it
typically conducts valuation. This section first describes the committee's understanding of the
Agency's needs and constraints related to ecological valuation. It then discusses the committee's
view of how ecological valuation is typically done at EPA, using an illustrative example. The
committee's observations from this section form the basis of its recommendations regarding use
of an expanded and integrated approach to valuation discussed in sections 4 and 5 of this chapter.

2.2.1. Policy Contexts at EPA Where Ecological Valuation Can be Important

As noted, much of the interest in ecological valuation at EPA stems from the need to

better represent the ecological benefits of EPA actions in analyses related to national rule
makings. Two of EPA's governing statutes (the Toxic Substances Control Act and the Federal
Insecticide, Fungicide and Rodenticide Act) require economic assessments for national rule
making. In addition, Executive Orders 12866 and 13422 have similar requirements for
"significant regulatory actions." A circular on "Regulatory Analysis" issued by the Office of
Management and Budget (OMB) in September 2003, OMB Circular A-4, identified key elements
of a regulatory analysis for "economically significant rules."7 . In developing the EBASP, EPA

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identified the need for improved models and methods to help implement the requirements of this
circular and provide better information on ecological effects that are currently not quantified or
monetized.

Valuation can also be useful to EPA in a second decision context, decision-making for
the remediation, restoration and redevelopment of contaminated sites. Decisions at clean-up
sites, whether they involve the hazardous waste sites listed on the Superfund National Priority
List that are eligible for federal cleanup funds under the Comprehensive Environmental
Response, Compensation, and Liability Act (CERCLA) or other clean-up sites (e.g., sites that are
the focus of EPA's Brownfields Economic Redevelopment Initiative, Federal Facilities
Restoration and Reuse Program, Underground Storage Tank Program, and Research
Conservation and Recovery Act), could be enhanced by ecological valuation that could
demonstrate the potential impact of ecological services obtained from site redevelopment.

A third decision context for valuation relates to EPA's regional office partnerships with
other governments and organizations where the contributions of ecological protection to human
welfare are potentially important. In this context, regional offices may find valuation useful in
priority setting, such as targeting projects for wetland restoration and enhancement or identifying
critical ecosystems or ecological resources for attention. Regional partnership efforts may also
involve assisting state and local governments, other federal agencies, and non-governmental
organizations with protecting lands and land uses. In these contexts, assessment of the value of
ecological protection options could aid in the decision making process and help partners
communicate the value of the option chosen.

Although many of the issues and recommendations throughout this report apply across
decision contexts, the committee recognizes that specific valuation needs and opportunities vary
across these contexts. For this reason, Chapter 6 of this report is devoted to detailed discussions
of the implementation of the report's general recommendations in these three specific decision
contexts: national rule making, site-specific restoration or redevelopment, and regional
partnerships. While the report discusses these three contexts explicitly, the committee also notes
that ecological valuation may also be useful for EPA in other contexts and for other purposes as
well. These include:

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•	Program assessments mandated by the Government Performance and Results Act
(GPRA) of 1993;8

•	Setting Supplemental Environmental Protection penalties for enforcement cases
where those penalties involve protection of ecological systems and services;

•	Choice of options for Superfund and Resource Conservation and Recovery Act
cleanups that could take ecological valuation into account;

•	Review of Environmental Impact Statements prepared by other federal agencies,
to comply with the National Environmental Protection Act; and

•	Executing ecological protection duties otherwise delegated to states, for those
specific states that have not applied for or been approved to run programs on their
own, such as issuing permits to protect water quality.

Although not discussed explicitly in this report, the committee believes that selected valuation
methods and the approach described in this report can be useful in the above contexts as well.

2.2.2. Institutional and Other Issues Affecting Valuation at EPA

The committee recognizes that EPA must conduct ecological valuation within a set of

institutional, legal, and practical constraints that affect what can be done to incorporate

ecosystem values into policy evaluations. These constraints include procedural requirements

relating to timing and oversight, as well as resource limitations (both monetary and personnel).

To better understand the implications of these issues for its work, the committee conducted a

series of interviews with Agency staff9 The interviews focused on the process of developing

economic analyses as part of Regulatory Impact Assessments (RIA) for rule making and on the

relationship between EPA and OMB. The interviews proved equally beneficial in leading to a

better understanding of strategic planning, performance reviews, regional analysis, and other

situations where the Agency has the need to assess the value of ecosystems and ecosystem

services.

EPA has a formal rule-development process involving several stages, each of which
imposes demands on the Agency. The Agency also develops rules to meet court-imposed
deadlines. Despite the rigidity of this process, there is no single way in which Agency analysts
assess the benefits of protecting ecosystems. Practices vary considerably across program offices,
reflecting differences in mission, in-house expertise, and other factors. Program offices have

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different statutory and strategic missions. The organization, financing, and skills of the program
offices differ enormously. The National Center for Environmental Economics (NCEE) is the
Agency's centralized reviewer of economic analysis within the Agency.10 Nonetheless, the
primary expertise and development of the rules resides within the program offices.

The timing of the process largely determines the kinds of analytical techniques that are
employed. This is related to court-imposed deadlines on the rule process, as well as to
intervening requirements related to the collection and analysis of new data. The scientific
community is accustomed to much longer time horizons for their analyses. Unfortunately,
collecting new data poses a significant bureaucratic problem for the Agency. To collect original
data, the Agency must submit an Information Collection Request, which is reviewed within the
Agency and by OMB. This hurdle is required by the Paperwork Reduction Act and imposes the
review responsibility on OMB, adding a significant amount of time to the assessment process.
With a year or two at most to conduct a study, this kind of review significantly limits the scope
of analysis the Agency can conduct. In particular, the Agency must by necessity rely heavily on
transferring both ecological and social values information from previous studies to the new
analysis.

Another issue is OMB's role in defining or directing ecosystem valuation exercises at
EPA. Among its activities, OMB acts as an oversight body that reviews EPA's economic benefit
analyses. EPA is required to provide sufficient justification for its claims regarding the
economic benefits of its actions, including any analyses of willingness to pay or willingness to
accept related to ecological protection. As noted above, EPA has been given explicit guidance
by OMB in the Circular A-4. For a contribution to human welfare or cost that cannot be
expressed in monetary terms, the circular instructs Agency staff to "try to measure it in terms of
its physical units," or, alternatively, to "describe the benefit or cost qualitatively" (p. 10).11
Thus, although Circular A-4 does not require that all economic benefits be monetized, it does
require, at a minimum, some scientific characterization of those contributions. Little guidance is
provided, however, on how to carry out this task. The circular instead urges regulators to
"exercise professional judgment in identifying the importance of non-quantified factors and
assess as best you can how they might change the ranking of alternatives based on estimated net
benefits" (p. 10).

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In conducting benefit assessments, EPA has an incentive to use methods that have been
accepted by OMB in the past. This creates a bias toward the status quo and a disincentive to
explore new or innovative approaches. The committee recognizes the importance of consistency
in the methods used for valuation, but also sees the limitations resulting from relying solely on
previously approved methods when innovative or expanded approaches might also be
considered.

A related issue involves review of RIAs by external experts. The Agency does not take a
standardized approach to RIA review. EPA staff and managers reported that peer review was
focused only on "novel" elements of an analysis, meeting the requirements of EPA's peer review
policy (U.S. Environmental Protection Agency, 2003; also see U.S. Environmental Protection
Agency 2006). This raises the question of how the Agency (and perhaps OMB) defines "novel."
Moreover, the novelty standard ironically creates a clear incentive to avoid conducting
innovative analyses since the fastest, cheapest option is to avoid review altogether.

Finally, the committee notes the importance of the organization of assessment science
within the Agency. The Agency relies, to varying degrees, upon a variety of offices to develop
assessments, including individual program offices and NCEE. It is not clear what form of
organization is most effective. A further complication is the availability and location of data
used to support ecological valuation. To resolve this issue, data that are housed within individual
program offices should be made public and readily shared with other offices.

The EBASP contains suggestions for addressing some of the limitations on ecological
valuation resulting from the Agency's internal structure. It advocates the creation of a high-level
Agency oversight committee and a staff-level ecological valuation assessment forum. The
committee endorses these efforts. Nonetheless, the Agency will continue to face significant
external constraints when considering ecological valuation. The committee recognizes the
practical importance of these constraints and urges the Agency to be as comprehensive as
possible in its analyses within the limitations imposed by these constraints.

2.2.3. An Illustrative Example of Economic Benefit Assessment Related to Ecological

Protection at EPA

To better understand the current state of ecosystem valuation at EPA, the committee
thoroughly examined one specific case in which assessment of economic benefits was
undertaken, namely, the environmental and economic benefits analysis that EPA prepared in

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support of new regulations for Concentrated Animal Feeding Operations (CAFOs) (U.S.
Environmental Protection Agency 2002a).12'13 In communications with the committee, the
Agency indicated that this analysis was illustrative in form and general content of other EPA
regulatory analyses and assessments of the economic benefits of ecological protection.

EPA proposed the new CAFO rule in December 2000, under the federal Clean Water
Act, to replace 25-year-old technology requirements and permit regulations. The final rule was
published in December 2003. The new CAFO regulations, which cover over 15,000 large CAFO
operations, require the reduction of manure and wastewater pollutants from feedlots and land
applications of manure and remove exemptions for stormwater-only discharges.

Because the proposed new CAFO rule constituted a "significant regulatory action" under
Executive Order 12866, EPA was required to assess the economic costs and benefits of the
rule.14 An intra-agency team at EPA, including economists and environmental scientists, worked
together with the U.S. Department of Agriculture on the economic benefit assessment. Prior to
publishing the draft CAFO rule in December 2000, EPA spent two years preparing an initial
assessment of the economic costs and benefits of the major options. After releasing the draft
rule, EPA spent another year collecting data, taking public comments, and preparing assessments
of new options. EPA published its final assessment in 2003. EPA estimates that it spent
approximately $1 million in overall contract support to develop the assessment, with
approximately $250,000-$300,000 allocated to water quality modeling.

EPA identified a wide variety of potential "use" and "non-use" benefits as part of its
analysis.15 Using various economic valuation methods, EPA provided monetary quantifications
in its CAFO report for seven benefit categories.16 Approximately eighty-five percent of the
monetary estimate of the benefits that were quantified by EPA was attributed to recreational
benefits. According to Agency staff, EPA's analysis was driven by what it could monetize.
EPA focused on those contributions for which data were known to be available for quantification
of both the baseline condition and the likely changes stemming from the proposed rule, and for
translation of those changes into monetary equivalents. EPA's final assessment provides only a
brief discussion of the contributions to human welfare that it could not monetize. The table in
the Executive Summary listed a variety of non-monetized contributions17 but designated them
only as "not monetized." EPA did not try to quantify these "contributions" in non-monetary
terms (e.g., using bio-physical metrics) or present a qualitative analysis of their importance.

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Instead, it represented the aggregate effect of these "substantial additional environmental
benefits" simply by attaching a "+B" place-holder to the estimated range of total monetized
benefits. Although the Executive Summary gives a brief description of these "non-monetized"
benefits, the remainder of the report devotes little attention to them.

Although it involved considerable effort, the CAFO economic benefits assessment
illustrates a number of limitations in the current state of ecosystem valuation at EPA. First, as
noted above, in implementing the Executive Order, the CAFO analysis did not provide the full
characterization of ecological contributions to human welfare using quantitative and qualitative
information, as required by the OMB Circular A-4. The report instead focused on a limited set
of economic benefits, driven primarily by the ability to monetize these benefits using generally
accepted models and existing value measures (transfer of economic benefits).18 These benefits
did not include all of the major ecological contributions to human welfare that the new CAFO
rule would likely generate, nor all of the contributions that generated public support for the new
rule.19 The Circular requires that an assessment identify and characterize all of the important
benefits of the proposed rule, not simply those that can be monetized. By focusing only on a
narrow set of contributions that could be readily monetized, the CAFO analysis and report
understate the total benefits of the rule change and distort the rationale supporting the final rule.
An unfortunate effect of this presentation is to suggest to readers that the benefits that were
monetized constitute the principal justification for the CAFO rule.20 In this case the focus on
monetized benefits did not affect the outcome of the regulatory review. It is certainly possible,
however, that in a different context an economic benefit assessment based only on easily
monetized benefits could inadvertently undermine support for a rule that would be justified
based on a more inclusive characterization of contributions to human welfare.

Second, the monetary values for many of the emphasized economic benefits were
estimated through highly leveraged benefit transfers that often were based on dated studies
conducted in contexts quite different from the CAFO rule application.21 This was undoubtedly
driven to a large extent by time, data, and resource constraints, which make it very difficult for
the Agency to conduct new surveys or studies and virtually force the Agency to develop benefit
assessments using existing value estimates. Nonetheless, reliance on dated studies in quite
different contexts raises questions about the credibility or validity of the benefit estimates. This

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is particularly true when values are presented as point estimates, without adequate recognition of
the underlying limitations due to uncertainty and data quality.

Third, EPA apparently did not embark on a comprehensive effort to model the rule's
ecological impacts. The report presents merely a simple conceptual model that traces outputs (a
list of pollutants in manure - Exhibit 2-2 in the CAFO report) through pathways (Exhibit 2-1) to
environmental and human health effects.22 This model provided useful guidance, but was not
sufficiently comprehensive to assure thorough analysis of the rule's ecological impacts. As a
consequence the analysis was unduly directed by Agency presumptions (or discoveries) about
the availability of relevant data and the likely opportunities to quantify effects precisely and to
link and monetize associated economic benefits. This was undoubtedly driven in part again by
the time pressures of putting together the regulatory impact analysis. Without a comprehensive
modeling effort at the outset, however, EPA had insufficient insight into the potential economic
benefits and other values that needed to be analyzed and estimated. Developing integrated
models of relevant ecosystems from the start of a valuation project would also help in identifying
important secondary effects, which frequently may be of even greater consequence or value than
the primary effects.23

Fourth, the CAFO analysis clearly demonstrates the challenges of conducting required
economic benefit assessments of ecological protection at the national level.24 National rule
making inevitably requires EPA to generalize away from geographic specifics, both in terms of
ecological impacts and associated values. It is, however, possible (and desirable) to make use of
existing and ongoing research at local and regional scales to conduct intensive case studies (e.g.,
individual watersheds, lakes, streams, estuaries) in support of the national-scale analyses. A key
question, of course, is whether case studies are representative. Both representative and non-
representative case studies can nonetheless provide useful information. Representative case
studies offer more detailed data and models that can fill in gaps in broad-scale national analyses
and check the validity of these analyses systematically. In general, systematically performing
and documenting comparisons to intensive study sites can indicate the extent to which the
national model needs to be adjusted for local or regional conditions. It also can provide data for
estimating the range of error and uncertainty in the projected national-scale effects. Non-
representative case studies can provide valuable information about the extent to which certain

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regions or conditions may yield impacts that vary considerably from the central tendency
predicted by the national analyses.

Fifth, although EPA invited public comment on the draft CAFO analysis as required by
Executive Order 12866, there is no indication in the draft CAFO report that the Agency
consulted with the public for help in identifying, assessing, and prioritizing the effects and values
addressed in its analysis. Nor is there discussion in the final CAFO analysis of any public
comments that might have been received on the draft CAFO analysis. Early public involvement
can play a valuable role in helping the Agency to identify all of the systems and services
impacted by the proposed regulations and to determine the regulatory effects that are likely to be
of greatest value. Through this added effort, valuations would be more likely to include the most
important impacts.

Sixth, EPA failed to follow its own advice regarding the use of outside peer-reviewed
data, methods, and models. While the Agency appropriately emphasized peer review in its
analysis and report, EPA did not seek peer review in deriving values for the CAFO rule. Once
again, this shortcoming is undoubtedly a function of time and resource constraints. It should be
noted, however, that peer review, especially early in the process, could help EPA staff identify
relevant and available data, models, and methods to support its analysis. In addition, it could
provide encouragement, direction, and sanction for more vigorous and effective pursuit of
ecological and human well-being effects associated with the proposed rule. An effective method
is to review not only individual components of the analysis (e.g., watershed modeling, air
dispersal, human health, recreation, and aesthetics) but the overall analytic scheme as well.

Finally, EPA's analysis and report closely adhered to the requirements of Executive
Order 12866, which provided the proximate reason for preparing the analysis and report.
Nevertheless, when EPA prepares a benefit assessment specifically to comply with Executive
Order 12866, the Agency need not limit itself to the goals and requirements of the Executive
Order, which directs EPA to conduct an "analysis" and "assessment" of the "benefits anticipated
from the regulatory action" and, "to the extent feasible, a quantification of those benefits." The
Executive Order, to be clear, does not preclude EPA from adopting broader goals. By adopting a
narrow focus, the CAFO report failed to consider the broader purposes served by a benefit
assessment. Assessments such as the CAFO study can serve many purposes, including helping
to educate policy makers and the public more generally about the economic benefits and other

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values that stem from EPA regulations. It is important for EPA to recognize this broader purpose
and to have an incentive to consider it more regularly.

2.3. An Integrated and Expanded Approach to Ecosystem Valuation: Key Features

The CAFO example discussed in the previous section highlights a number of limitations

to the current state of ecosystem valuation at EPA. The committee's analysis points to the need
for an expanded, integrated approach to valuing the ecological impacts of EPA actions, focusing
on the impacts of greatest concern to people and integrating ecological analysis with valuation.
This section describes an approach to ecological valuation developed and endorsed by the
committee. The approach should serve as a guide to EPA staff as they conduct RIAs and seek to
implement the provisions of Circular A-4, as well as in decisions regarding regional and local
priorities and activities. A more detailed discussion of the implementation of the approach and
the framework for specific decision contexts is provided in subsequent chapters of this report.

As noted, the committee focused on valuation in EPA contexts where there is an
environmental protection decision to be made. The major components of the ecological
valuation process proposed by the committee are depicted in Figure 1.

The committee's proposed approach for implementing the valuation process has three
key, interrelated features: a) early consideration of effects that are socially important; (b)

Figure 1: Components of Ecological Valuation

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predicting ecological changes in value-relevant terms; and (c) drawing on a suite of methods for
characterizing values.

2.3.1. Early Consideration of Effects that are Socially Important

The first key component of the proposed approach is the early identification and

prediction of the impacts or contributions to human welfare that are likely to be most significant
or of greatest importance to people, whether or not the impacts are easily measured, monetized
or widely recognized by the public. These could include changes in the ecosystem itself that
people value directly, or the resulting changes in the ecological services provided by a system.
Information about the ecosystem services or characteristics that are of greatest concern needs to
be obtained early in the valuation process so that efforts to quantify and characterize values can
be focused on the related ecological changes. The importance of a given change will depend on
the magnitude and bio-physical importance of the effect and on the resulting importance to
society.

Identifying socially relevant effects requires a systematic consideration of the many
possible sources of value from ecosystem protection and an identification of the types of values
that provide the impetus for a particular policy change. This focus will likely lead to an
expansion of the types of services to be characterized, quantified, or explicitly valued. For
example, even in the context of national rule making, a specific contribution to human welfare
should be included as part of an overall valuation whether or not it is possible to monetize that
benefit in terms of willingness to pay or willingness to accept; if there is evidence that it is
important to people, the benefit should be included as a key component of the total benefits,
complete with a detailed and careful (even if not monetized) characterization of its importance.
Previous assessments have often focused on what can be measured relatively easily rather than
what is most important to society. This diminishes the relevance, usefulness, and impact of the
assessment.

An obvious question is how to assess the likely importance of different ecological
impacts prior to completion of the valuation process. In fact, a main purpose of conducting a
thorough valuation study is to provide an assessment of this importance. Nonetheless, in the
early stages of the process, preliminary indicators of likely importance can be used as screening
devices to provide guidance on the types of impacts that are likely to be of greatest concern.
Relevant information can be obtained in a variety of ways. Examples range from in-depth

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studies of people's mental models and how their preferences are shaped by their
conceptualization of ecosystems and ecological services, to more standard survey responses from
prior or purpose-specific studies. In addition, early public involvement25 or the use of focus
groups or workshops, comprised of representative individuals from the affected population and
relevant scientific experts, can help to identify ecological changes for the specific context of
interest.

In eliciting information about what matters to people, it is important to bear in mind that
people's preferences depend on their mental models (i.e., their understandings of causal
processes and relations), the information that is at hand to influence their understanding, and how
that influence occurs. As noted previously, expressions of what is important (e.g., in surveys) or
of the tradeoffs people are willing to make can change with the amount, the manner and the kind
of information provided. Collaborative interaction between analysts and public representatives
can help to ensure that respondents have sufficient information when expressing views and
preferences. The ecological valuation process can in fact be used as a mechanism for educating
the public about the services provided by ecological systems and how those services are affected
by EPA actions, thereby narrowing the gap between expert and public knowledge of ecological
effects.

2.3.2. Predicting Ecological Changes in Value-relevant Terms

The second major component of the C-VPESS process is the need to predict ecological

changes in terms that are relevant for valuation. This requires both the prediction of bio-physical
impacts of EPA actions using ecological models and the mapping of those impacts into changes
in ecosystem services or features that are of direct concern to people. Ideally, this would be done
using an ecological production function that is specified and parameterized for the ecosystem
and associated services of relevance in the assessment.

Numerous mathematical models of ecological processes and functions have been
developed. These models cover the spectrum of biological organization and ecological hierarchy
(e.g., individual level, the population level, the community level, the ecosystem level, and the
level of the global biosphere). They can be used to predict ecological impacts associated with a
given EPA action at different temporal and spatial levels. Some have been developed for
specific contexts, such as particular species or geographic locations, while others are more
general.

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Ecological models provide a basis for estimation of the ecological changes that could
result from a given EPA action or policy (e.g., changes in net primary productivity or tree
growth, bird or fish assemblages) and the associated changes in ecosystems or ecosystem
services. However, many of these ecological models have been developed to satisfy research
objectives and not EPA policy or regulatory objectives. This poses challenges when using these
models to assess the contributions of EPA actions to human welfare.

The first challenge is to link existing models with Agency actions that are intended to
control chemical, physical, and biological sources of stress. The valuation framework outlined
here requires an estimation of the bio-physical impacts that would stem from a specific EPA
action. To be used for this purpose, ecological models must be linked to information about
stressors. This link is often not a key feature of ecological models developed for research
purposes.

Ecological models additionally need to be appropriately parameterized for use in policy
analysis. Numerous detailed ecological studies have been conducted at various levels, for
example, at Long-Term Ecological Research Sites (Farber et al. 2006). These could provide a
starting point for parameterizing policy-relevant models. A key challenge is to determine
whether (or to what extent) parameters estimated from a given study site or population can be
"transferred" for use in evaluating ecological changes at a different location, time, or scale. In
other words, to what extent are estimated parameters adaptable from one context to another in
estimating the contributions to human welfare and values associated with EPA actions? In many
cases, data do not currently exist to parameterize existing models so that they could be used in
assessing EPA's actions. Such data may need to be developed before the Agency can use these
models fully. To the extent that transferable models and parameter estimates exist, it would be
extremely valuable to have a central depository that EPA could draw on for this information.

The final, but perhaps most important, challenge is translating the changes predicted by
standard ecological models into changes in ecosystem services or features that can then be
valued. If adapted properly, ecological models can connect material outputs to stocks and
services flows (assuming that the services have been well-identified). Providing the link
between material outputs and services involves several steps, including identifying service
providers, determining the aspects of ecological community structure that influence function,
assessing the key environmental factors that influence the provision of services, and measuring

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the spatial and temporal scales over which services are provided (Kremen 2005). Most
ecological models, however, are not currently designed with this objective in mind. In particular,
they do not translate bio-physical impacts in ways that lay individuals can understand, or in ways
that reflect how those changes are of value to them.

2.3.3. Drawing on Multiple Methods for Characterizing Values

Given predicted ecological changes, the value of these changes needs to be characterized

and, when possible, measured or quantified. As noted above, a variety of valuation methods
exist. The C-VPESS approach envisions drawing on a wider range of methods than EPA has
typically utilized to capture a broader array of values. It recognizes that there are many sources
and types of value and many valuation methods. In addition, different methods provide different
ways of characterizing information about values, and multiple methods may be needed to
sufficiently capture all types or sources of value. Given the array of values and methods, a key
tenet of the valuation process proposed by the committee is that each valuation process should
include a conscious choice regarding the type(s) of value to assess and the appropriate methods
for assessing those values. However, this expanded approach should include only those methods
that meet accepted scientific standards of precision and reliability, are appropriately responsive
to relevant changes in ecosystems and their services, and are properly related conceptually and
empirically to things people value. The suite of methods used could vary with the specific policy
context, due to differences across contexts in information needs, legal and regulatory
requirements, the underlying sources of value being captured, data availability, and
methodological limitations.

Through expanded methodology EPA can better capture the full range of contributions
stemming from ecosystem protection and the multiple sources of value derived from ecosystems.
In addition, where resources allow, the use of multiple methods to characterize the same
underlying value can in some cases increase the confidence that decision makers, policy makers,
and the public have in those estimates. Certainly, the possibility exists that the application of
multiple assessment methods to an environmental decision problem could suggest conflicting
information about relative values. It then would be essential to try to ascertain the source of the
differences. In some cases, they may be due to the application of methodologies (e.g., eliciting
values from different population groups or samples), or study limitations (e.g., inappropriate
application of techniques or interpretation of results), or the inherent uncertainty in estimating

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values that results from data limitations, theory limitations, and randomness (see related
discussion in Chapter 5). In other cases the differences may reflect the fact that the alternative
methods are capturing fundamentally different sources, components, or concepts of value. In
any case, information regarding the similarities or differences of alternate assessment methods,
including their conclusions about the value of an ecological change, would be an important input
into a policy decision.

The committee evaluated a number of different methods for characterizing values
(described in detail in Chapter 4 and Appendix B). These include not only economic valuation
methods (the usual focus of EPA valuation) but other methods that could be used to value
ecological changes as well. These include social and psychological methods, assessments based
on voting and other group expressions of social or civic values, and assessment methods based
on indicators or bio-physical rankings that are less directly dependent on human preferences and
value judgments.

Underlying many valuation methods (including preference-based methods) are metrics
that are primarily bio-physical or socio-economic indicators of impact. These include such
indicators as acres of habitat restored, the number and characteristics of individuals or
communities affected, the likely injuries avoided, and the duration of impact. These metrics can
provide useful information in at least three ways. First, in some cases, they can be used directly
in policy decisions. For example, decisions based on human impact criteria (e.g., protection of
children's health) or environmental goals (such as promotion of biodiversity) may draw directly
from these measures as indicators of the appropriate policy choice. Second, they might be used
as a proxy for some component of the contributions of ecosystem protection to human welfare,
when that component cannot be readily valued. As noted earlier, in contexts requiring benefit-
cost analyses, the OMB Circular A-4 requires that benefits be quantified when they cannot be
monetized; these metrics provide potentially useful forms of quantification in such
circumstances. Finally, even when human impacts can be valued, these metrics provide
information about human impacts that would presumably be relevant in the determination of the
associated value of the ecological change. Thus, in all of these contexts, estimates of the impact
of the ecosystem change on human populations are needed.

In contexts where monetary metrics are required or desired and the necessary data and
methods exist, the impact of the ecological change on the provision of some services to human

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populations may be translated into a monetary equivalent of that change using standard
economic valuation techniques. For some valuation contexts, economic methods for valuing
changes are relatively well developed. As noted previously, existing EPA ecological valuation
efforts, such as the EBASP and the Science to Achieve Results (STAR) Grant program, have
focused on valuing changes using economic methods. These methods are designed to estimate
the economic benefit or cost of a given ecological change using a willingness-to-pay or
willingness-to-accept measure of the utility equivalent of that change. They have been applied to
the valuation of ecosystem services in a number of studies that have produced results that are
useful for policy evaluation and decision making.

As in the CAFO study, however, economic valuation methods have generally been
applied to a relatively narrow set of services. In some cases, these services might not have been
those that people are most concerned about protecting. While there are continuing discussions
about the role of economic valuations in principle, it is unlikely as a practical matter that all of
the important benefits (or costs) of a change in ecological conditions will be sufficiently captured
by economic valuation methods. For this reason, the EBASP calls for exploring "supplemental"
approaches to valuation.

The valuation approach proposed by this committee calls for a more prominent role to be
played by a variety of methods for characterizing values than envisioned in the EBASP. This is
a practical alternative for use when economic methods cannot fully capture contributions to
human welfare because of limitations in data or other knowledge-based gauges. It is also a means
of capturing many components of value that are not fully reflected in those value measures that
are based solely on economic measures of willingness to pay or willingness to accept. Including
other scientifically-based assessment approaches that can be applied along with, or in place of,
economic assessments will allow EPA to more fully represent the contributions of ecosystems
and their services to human well-being.

2.4. Steps in Implementing the Proposed Approach

The previous subsections provide an overview of an integrated and expanded approach to

ecological valuation proposed by the committee. The process for implementing the proposed
framework would involve the following steps, depicted in Figure 2:

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1.	Formulating the valuation problem and choosing policy options to be considered,
given the policy context;

2.	Identifying the significant bio-physical changes that could result under the
different options;

3.	Identifying the changes in the ecosystem and its services that are socially
important;

4.	Predicting the changes in the ecosystem and relevant ecosystem services in
biophysical terms;

5.	Characterizing, representing, or measuring the value of changes in the ecosystem
and its relevant services in monetary or non-monetary terms; and

6.	Communicating results to policymakers for use in policy decisions.

Although Figure 2 depicts these steps as sequential, in practice interactions and iterations across
steps are likely during the process. For example, information about the value of changes in
ecosystem services stemming from a given set of policy options might cause a reformulation of
the problem or identification of new policy options that could be considered. Also, a projected
bio-physical effect might suggest human-social values that were not captured in initial
public/stakeholder processes.

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Figure 2: Process for Implementing an Expanded and Integrated Approach to Ecological Valuation

As depicted in Figure 2, the implementation of the approach is contingent upon the
specific policy context. As noted above, ecological valuation can play a key role in a number of
different decision contexts, including national rule making and regional or local decisions
regarding priorities and actions. The valuation problem should be formulated within the specific
EPA context. Different contexts will generally be governed by different laws, principles,
mandates, and public concerns. These contexts can differ not only in the required scale for the
analysis (e.g., national vs. local) but possibly also in the type of valuation information that is
needed. For example, in contexts requiring an economic benefit cost analysis, benefits need to
be monetized whenever possible. In contrast, expressing contributions to human welfare in
monetary terms might be of little or no relevance to EPA analysts in other contexts. The policy
context in which the assessment is cast is therefore a key influence on the appropriateness of
data, models and methods.

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Figure 2 also highlights the need for information and input from a wide range of
disciplines at each step of the process, beginning with problem formulation and the identification
of the impacts that matter to the estimation of the value of those impacts. Thus, instead of having
ecologists work independently from economists or other social scientists, this approach envisions
collaborative work across disciplines. The result is an analysis that identifies the impacts that are
of greatest concern to society in a manner that is informative for valuation. Ecological models
need to be developed, modified, or extended to provide usable inputs for value assessments.
Likewise, valuation methods and models need to be developed, modified, or extended to address
important ecological/bio-physical effects that are currently underrepresented in value
assessments.

Figure 2 additionally suggests a structure that in many ways parallels the Agency's
Framework for Ecological Risk Assessment (U.S. Environmental Protection Agency Risk
Assessment Forum 1992; U.S. Environmental Protection Agency Risk Assessment Forum 1998).
This framework underlies the ecological risk guidelines developed by EPA to support decision
making that is intended to protect ecological resources (U.S. Environmental Protection Agency
Risk Assessment Forum 1992). The committee views ecological valuation as a complement to
ecological risk assessment. Both processes begin with an EPA decision or policy context
requiring information about ecological effects. Following that is a formulation of the problem
and an identification of the purpose and objectives of the analysis as well as the policy options
that will be considered. In addition, both ecological risk assessment and ecological valuation
involve the prediction and estimation of possible ecological effects of an EPA action or decision
under consideration. They also both ultimately use this (and related) information in the
evaluation of alternative decisions or policy options.

Ecological valuation goes beyond ecological risk assessment in an important way. Risk
assessments typically focus on predicting the magnitudes and likelihoods of possible adverse
effects on species, populations, and locations, but do not provide information about the societal
importance or significance of these effects. In contrast, as depicted in both Figure 1 and Figure
2, ecological valuation seeks to characterize the importance to society of predicted ecological
effects by providing information on the value that society places on either the ecological
improvements or the loss it experiences from ecological degradation. By incorporating human
values, ecological valuation is closer to risk characterization than risk assessment, and many of

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the principles that should govern risk characterization outlined in the 1996 NRC Report
Understanding Risk: Informing Decisions in a Democratic Society would pertain to ecological
valuation as well. For example, both should be the outcome of an analytical and transparent
process that incorporates not only scientific information but also information from the various
interested and affected parties about their concerns and values.

In contexts involving complex ecological impacts and tradeoffs, deliberative processes
have been successfully used as a means of identifying stakeholder concerns, educating
stakeholders about the ecological impacts of alternative policy choices, eliciting information
about stakeholder values, and ultimately describing and possibly evaluating tradeoffs. Examples
include the decision-aiding processes developed by decision scientists (refs) and mediated
modeling, in which stakeholders participate in the development and interactive use of simulation
models of complex ecological systems to compare and evaluate policy options (refs). The
process in Figure 2 has a structure that parallels these deliberative processes and shares many of
the same goals. In some contexts (e.g., site-specific and regional valuations), a single, holistic
deliberative process could be applied in a very similar way to accomplish the entire valuation
process. In other contexts, implementation of the valuation process could involve elements of a
deliberative process at different points in the overall value assessment (e.g., early on when
identifying impacts that are socially important or educating the stakeholders about potential
impacts), coupled with the use of non-deliberative methods at other stages of the process. In
either case, the goals and overall structure of this report's proposed valuation approach closely
parallel those of the deliberative processes that have been developed and successfully used in a
number of contexts.

2.5. Conclusions and Recommendations

Ecosystems play a crucial role in supporting life as we know it. They provide a wide

array of services that directly or indirectly support or enhance human populations. In addition,
they can be valued in their own right, for non-anthropocentric reasons stemming from ethical,
religious, cultural or biocentric principles. Part of EPA's broad mission to protect human health
and the environment includes the protection of ecosystems.

Many EPA actions affect the state of ecosystems and the services derived from them.
However, to date ecosystem impacts have received relatively limited consideration in EPA
policy analysis, which has typically focused on human health impacts. It is imperative that EPA

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improve its ability to value ecosystems and their services to ensure that ecological impacts are
adequately considered in addition to human health impacts in the evaluation of EPA actions at
the national, regional and local levels.

To date, ecological valuation at EPA has focused primarily on a limited set of
contributions to human well-being from ecological protection. This stems primarily from the
difficulty of predicting the impact of EPA actions on ecological systems and the services derived
from them and the difficulty of quantifying, measuring, or characterizing the resulting
contributions and associated values. The presumption that contributions need to be monetized in
order to be carefully characterized also restricts the range of ecological impacts that are typically
considered in EPA analyses, particularly at the national level.

The committee views EPA's efforts to improve its ability to value ecological systems and
services as very important and timely. The committee recommends that the Agency move
toward covering an expanded range of important ecological effects and human considerations
using an integrated approach. Such an approach would:

a)	Expand the range of ecological changes that are valued, focusing on valuing the
ecological changes in systems and services that are most important to people and
recognizing the many sources of value, including both instrumental and intrinsic
values;

b)	Highlight the concept of ecosystem services and provide a mapping from changes
in ecological systems to changes in services or ecosystem components that can be
directly valued by the public; and

c)	Utilize an expanded set of methods for identifying, characterizing, and measuring
the values associated with these changes.

Such an approach would, from the beginning and throughout, involve an interdisciplinary
collaboration among physical/biological and social scientists, as well as direct and early
involvement and input from the public or representatives of individuals affected by the
ecological changes. In implementing the approach, EPA should recognize the multi-dimensional
nature of value and make a conscious choice regarding the type of value(s) it wants to assess and
the appropriate methods for assessing those values. In addition, the Agency should be

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transparent about the reasons for choosing specific valuation methods and communicate clearly
what the methods that it chooses measure and do not measure.

Through the use of the expanded and integrated valuation framework recommended in
this report, EPA can move toward greater recognition and consideration of the effects that its
actions have on ecosystems and the services they provide. This will allow EPA to improve
environmental decision-making at the national, regional and site-specific levels and contribute to
EPA's overall mission regarding ecosystem protection. In addition, EPA can better use the
ecological valuation process as a mechanism for educating the public about the role of
ecosystems and the value of ecosystem protection. The remainder of this report develops the
ideas embodied in the C-VPESS integrated value assessment system through a more detailed
look at how the approach could be applied.

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3 BUILDING A FOUNDATION FOR ECOLOGICAL VALUATION:
PREDICTING EFFECTS ON ECOLOGICAL SYSTEMS AND

SERVICES

Chapter 2 of this report presented an overview of an integrated and expanded
approach to valuing ecological changes that result from EPA actions or decisions. The
approach was described in general terms. This chapter focuses on one part of that approach,
namely, predicting ecological changes in value-relevant terms. No matter what valuation
method is used, the valuation process requires an assessment of the impact of a given EPA
action on ecosystems and the services they provide. To conduct the assessment, a prediction
of the bio-physical impacts is needed in terms that are relevant for ecological valuation. To
the extent possible, this prediction should be quantitative. In the context of national rule
making, quantification is necessary for values that will be monetized and is needed (as stated
in Circular A-4 from the Office of Management and Budget), even for values that cannot be
readily monetized. In every context where the need for valuation arises, information about
the magnitude of effects will be a key component of value assessment.

This chapter begins with a discussion of the importance of developing an initial
conceptual model of the relevant ecosystem and its services designed to guide the entire
valuation process. It then turns to a discussion of how to operationalize the conceptual
model, which will often involve the use of multiple specific ecological models. In this
context, the key role played by the concept of an ecological production function is discussed.
The discussion highlights the challenges that currently exist in trying to implement ecological
production functions in specific valuation contexts. These include challenges associated with
understanding and modeling the relevant ecology, clearly identifying the relevant ecosystem
services, and mapping ecological changes into changes in the ecosystem services of interest.
To a large extent, these challenges stem from the underlying complexity and site-specificity
of ecosystems. The chapter then discusses some strategies for addressing these challenges
and providing the ecological science necessary to support valuation. A final section
summarizes conclusions and recommendations.

3.1. The Road Map: A Conceptual Model

Formulation of a conceptual model is a key first step in predicting the effects of EPA

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actions on ecological systems and services. The committee recommends that EPA start each

ecological valuation by developing a conceptual model of the relevant ecosystem(s) and

associated services. The conceptual model should be constructed at a general level to

provide a road map to guide the valuation process. As a result, the model should be context-

specific. More detailed analyses involving ecological production functions should follow to

identify the key interactions, predict specific ecological impacts, and calculate the ecological

values. This will often require the use of ecological or valuation-related models with a

narrower focus (see section 3.3). The conceptual model's basic purpose is to guide the

process by providing a framework for integrating these more specific analyses into the

overall valuation exercise.

Key features of the conceptual model are a clear identification of the relevant
functional levels of the ecosystem, the inter-relationships between ecosystem components,
and how they contribute to the provision of ecosystem services, either directly or indirectly.
An example illustrating some aspects of ecosystem services related to nutrient pollution is
provided in Figure 3, adapted from Covich et al. (2004).

Figure 3: Illustration from Covich et al., 2004, Showing Relationships of Major Functional Types to

Ecological Services

FOOD SUPPLY

4

macroinvertebrate
grazers

phytobenthos,

biofilms,
aquatic plants

CLEAN WATER	RECREATION

(drinking water, irrigation) (e.g., hunting, fishing, boating)

4

fish, waterfowl, and other aquatic vertebrates

i

i



i

J





invertebrate

predators



f

t

	 t



filter/deposit feeders

shredders

N

fragmentation j

leakage
FPOM	|

protozoans
flagellates

phyto- and
bacterio plankton

dissolved nutrients



WASTE
DISPOSAL

TERRESTRIAL INPUT
OF ORGANIC MATTER

NUTRIENT LOADING
;.g„ weathering of soils)

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Figure 3 highlights the need for the conceptual model to include both information
about the underlying ecology and a link to ecological services that are of importance to
society. There is a need to include, for example, the impacts of environmental stressors, such
as waste disposal on organisms at different trophic levels, the key interactions among species
at different levels, and the changes at different levels that affect ecological services, such as
the food supply, clean water, or recreation.

Ecologists, not surprisingly, often focus on the underlying ecological aspects
(depicted in the lower part of Figure 3), while valuation experts tend to focus on the later,
value-oriented stages of the process, starting with ecosystem services (i.e., starting at the top
of Figure 3). A key principle of the C-VPESS approach is the need to consider and integrate
both aspects of the process. For ecological valuation aimed at improved decision-making, a
detailed analysis of ecological impacts, including modeling of ecosystem impacts, is
insufficient unless those impacts are mapped to changes in ecological services or system
components of importance to people. It is similarly insufficient to conduct valuation
exercises that do not reflect the key ecological processes and functions affected by the
decisions under consideration. Both steps are essential, and the development of a conceptual
model at the outset of the valuation process can help ensure that the process is guided by this
basic principle.

As envisioned here, the development of the conceptual model is a significant task that
deserves the attention of EPA staff throughout the agency, experts in the relevant topics of
consideration (from both the bio-physical and social sciences), and the public. Involving all
constituents at this stage will enhance transparency, provide the opportunity for more input
and better understanding, and ultimately give the process more legitimacy. Participatory
methods such as mediated modeling (see Appendix B) can play a valuable role in the
development of the conceptual model.

In addition, the process for development of the conceptual model should allow for

iteration and possible model changes and refinement over time. For example, an action at a

local site may initially be considered to have only local ecological effects, but, once the

stressors are considered, it may become apparent that effects reach to more distant regions

downstream or down wind, requiring a change in the conceptual model. Similarly, as the

stressors are identified in the context of the relevant ecological system, the conceptual model

may need to be modified to incorporate additional stressors. As an example, a relatively non-

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toxic chemical effluent, normally seen as insignificant, might become significant if it is

determined that low stream flows or intermittent streams effectively increase the

concentration of the chemical to toxic levels during some parts of the year. The conceptual

model, the process for developing and completing it, and the decisions that are embedded

within it should all be a part of the formal record.

3.2. Operationalizing the Conceptual Model: The Role of Ecological Production
Functions

While the conceptual model serves as a guide for the overall valuation process in a

specific context, the individual components and linkages embodied in that model must be

operationalized. The goal is to provide, to the extent possible, quantitative estimates of the

changes in ecosystem components or services that can then be valued. To operationalize the

conceptual model, it is necessary to map or describe: a) how the EPA action will affect the

ecosystem, b) how the change in the ecosystem will lead to a change in the provision of

ecosystem services; and c) how people value that change in ecosystem services. For the first

step, it is necessary to describe how change in stressor or in some other environmental factor

that could be altered by the EPA action results in changes in important aspects of ecosystem

structure or function. Does the change in stressor cause a species to disappear or change in

abundance? Does it result in a change in biogeochemistry? For any important changes, a

quantitative relationship must be determined.

A fundamental concept for describing the second step in this mapping is the

ecological production function. Ecological production functions are similar to the production

functions used in economics to define the relationship between inputs (labor, capital

equipment, raw materials) and outputs of goods and services. For example, a farmer uses

inputs of seeds, fertilizer, labor, and equipment to produce outputs of agricultural crops.26

Ecological production functions describe the relationships between ecological inputs

and outputs, i.e., between the structure and function of ecosystems and the provision of

various ecosystem services. These functions capture the biophysical relationships between

ecological systems and the services they provide, as well as the inter-related processes and

functions, such as sequestration, predation, and nutrient cycling. Expanding on the farming

example, in addition to the inputs mentioned above, there are ecological inputs provided by

ecosystems, such as soil nutrients, rainfall, and pollinators, that have a major impact of crop

production. However, crop production is not the only ecosystem service provided by these

inputs. Beyond crop production, additional important outputs (i.e., ecosystem services)

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provided by agriculture include the effect of farming operations on carbon sequestration,

water quality, habitat of pollinators and other species. An ecological production function

could be developed for each of these services separately. Alternatively, to the extent that

some services are linked (e.g., produced jointly or in competition), a multiple-output function

could be developed to capture these linkages.

Thus, ecological production functions generate an accounting of the relationship
between a broad suite of inputs and a broad suite of goods and services from ecosystems.
Coupled with information about how changes in stressors affect the ecological inputs, these
functions can be used to predict the changes in ecosystem services that will result from
alternative Agency actions or management scenarios. In addition, they allow answers to
questions such as: How can forests be managed to reduce catastrophic damage from fire?
What kinds of marine reserves lead to larger fish populations? How much more wetland is
needed to recharge sub-surface aquifers used for irrigation?

Implementing the concept of an ecological production function requires: a)
characterization of the ecology of the system, b) identification of the ecosystem services of
interest; and c) development of a complete mapping from the structure and function of the
ecological system to the provision of the relevant ecosystem services. Figure 4 provides a
graphical representation of how this concept can be implemented. The left-hand side
represents ecological models at various organizational levels that are used to predict
ecological endpoints (see further discussion of endpoints below). While these are important
components of an ecological production function, it is not the complete function. An
ecological production function requires that the predictions regarding the levels or changes in
these ecological endpoints be translated into corresponding predictions regarding ecosystem
services.

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Figure 4: Graphical Depiction of Ecological Production Functions

ECOLOGICAL PRODUCTION FUNCTIONS

Ecological Models Ecological Endpoints Map to Ecological Services

Rapid progress is being made in understanding ecological production functions for
certain ecosystem services. One such service is pollination. Animal pollination is essential
for the production of about one-third of agricultural crops and the majority of plant species
(Kremens et al. 2007). Ecologists have recently built spatially explicit models incorporating
land use and its effect on habitat and foraging behavior of pollinators (Kremens et al. 2007).
One application of such models is to link changes in ecosystem conditions to the level of
pollination of agricultural crops and their yields. Empirical studies using this approach have
shown the effects of proximity to natural forest on coffee productivity (Ricketts et al. 2004)
and the interaction of wild and honey bees on sunflower pollination (Greenleaf and Kremens
2006).

A second ecosystem service where considerable progress has been made in
developing ecological production functions is carbon sequestration. Agricultural systems,
forests and other ecosystems contain carbon in soil, roots, and above ground biomass.
Rapidly growing markets for carbon and the potential for generating carbon credits are

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pushing interest in accurately assessing the carbon sequestration potential of agricultural and

other managed ecosystems (Willey and Chamaides 2007). It is possible to quantify above

ground carbon stores in various types of ecosystems such as forests fairly accurately (e.g.,

Birdsey 2006, Smith et al. 2006, U.S. Environmental Protection Agency Office of

Atmospheric Programs 2005), while greater uncertainty remains about stocks of soil carbon

that make up the majority of carbon in agricultural and grassland systems (e.g., Antle et al.

2002, U.S. Environmental Protection Agency Office of Atmospheric Programs 2005).

Despite this progress, our current understanding of ecological production functions
for the complete range of services from ecosystems remains limited (Balmford et al. 2002,
Millennium Ecosystem Assessment 2005, National Research Council 2004). Although many
ecological models exist (see further discussion below), most of these are not designed to link
changes in ecological inputs or endpoints to changes in ecosystem services. The following
section discusses some of the challenges in developing ecological production function
models for use in ecological valuation.

3.3. Challenges in Implementing Ecological Production Functions

3.3.1. Understanding and Modeling the Underlying Ecology

As noted above, operationalizing the conceptual model using ecological production

functions requires a fundamental understanding of the components, processes, and
functioning of the ecosystem(s) that underlie and generate the ecosystem services. In other
words, analysts must have a strong understanding of the underlying ecology. While much is
known about ecological systems, current knowledge is still very incomplete due in large part
to the fact that ecosystems are inherently complex, dynamic systems that vary greatly over
time and space.

As an example of the complexity of ecological functions, consider the ecological
services associated with the activities of soil organisms that might be affected by disposal of
waste on that soil. These organisms thrive on organic matter present or added to the soil. By
breaking down that organic matter, certain groups of organisms maintain soil structure
through their burrowing activities, which in turn provide pathways for the movement of
water and air. Other kinds of organisms shred the organic material into smaller units that are
in turn utilized by microbes. These microbes then release nutrients in a form that can be
utilized by higher plants for their growth or in a dissolved form that is hydrologically
transported from the immediate site into the water table or stream. Other groups of

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specialized microbes may release various nitrogen gases directly to the atmosphere. Thus,

the nature of the soil organisms and the products that they utilize, store, or release all help to

regulate the biogeochemistry of the site as well as its hydrology, productivity, and carbon

storage capacity. As Figure 4 suggests, these kinds of functions relate to the services that

people more readily appreciate and value, such as the natural processing of wastes and the

provision of clean water (Wall 2004). This evaluation requires an understanding of the

complex ecological relationships that contribute to these services.

Complexity also stems from the fact that ecological effects may persist for different
periods of time (e.g., carbon dioxide in the atmosphere vs. acute toxic exposures to hazardous
chemicals), affecting both the temporal and spatial scales that are relevant for any analysis.
Numerous studies including EPA's regional analyses, risk analyses and the Environmental
Monitoring and Assessment Program (EMAP) provide guidance in identifying the proper
boundaries and time scales for the ecological system under study as well as the ecosystem
characteristics, stressors and endpoints (Harwell, et al., 1999, Young and Sanzone, 2002).

Because of the complexity of most ecosystems, models are used to organize
information, elicit the interactions among the variables represented in the models, and reveal
outcomes when run under different sets of assumptions or driving variables.27 Ecological
models can describe ecological systems and ecological relationships that range in scale from
local (individual plants) to regional (crop productivity) to national (continental migration of
large animals). As shown in Figure 4, these models frequently focus on specific ecological
characteristics, such as populations of one or more species or the movement of nutrients
through ecosystems, and can cover the spectrum of biological organization and ecological
hierarchy. For instance, a hydrological model might describe possible changes in the timing
and amount of water in streams and rivers. A biogeochemical model could predict effects on
the levels of various chemical elements in soils, ground water, and surface waters. A
terrestrial carbon cycle model could project changes in plant growth and in carbon sinks or
sources. Population and community models would project changes in specific animal and
plant populations that are of concern.

Primers on ecological theory and modeling such as Primer of Ecological Theory
(Roughgarden 1998b) can provide a starting point for identifying available models. Some
models are statistical, while others are primarily simulation models. Some statistical and
theoretical models are relatively small, containing a few equations. Other ecological models

are very large, involving hundreds of interacting calculations.

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Although many ecological models are well established and used routinely for

describing ecological systems, ecological models can only represent the current state of
knowledge about the dynamics of an ecological system and generate outputs as reliable as the
data the models use. The dynamism of a system adds to the challenge of modeling, as does
the non-linear responses of system components. The model outputs are estimates with
known, or sometimes unknown, levels of statistical uncertainty. No ecological model can
include all possible interactions. Some ecological models explicitly or implicitly incorporate
human dimensions, but most focus primarily on ecological functions. Models additionally
capture historical relationships and typically are not able to predict ecosystem patterns for
which no modern counterpart exists. For example, if a stressor such as climate change can
lead species to "reshuffle into novel ecosystems unknown today" for which there are no
analog, current models will not predict theses impact (Fox 2007).

Finally, the applicability - and to some degree the formulation - of ecological models,
is frequently constrained by the insufficiency of data to build and test the models. Even
when a full theoretical model of the ecosystem exists, that model will need to be
parameterized for the specific valuation context of interest. However, parameterization is
generally difficult because of the complexity of ecological systems and their dependence on
an array of site-specific variables. Many ecological models, as a result, are site specific.
Moreover, the relatively large amounts of site-specific data required to build and
parameterize models means that their transferability is limited, either because the model has
been developed using spatially constrained data or because inadequate data are available at
secondary sites with which to drive or parameterize the model. This site-specificity may
significantly limit the models' applicability to the spatial and temporal complexities required
in valuing ecological services, especially at regional and national scales.

Despite these caveats, utilizing ecological models provides a means of incorporating

the best available scientific knowledge of how ecosystems will respond to a given

perturbation and the sensitivity of various ecosystem components. Hence, they provide an

essential way to represent and ecological production functions and allow them to be

analyzed. Guided by the conceptual model, ecological models should be utilized to quantify

the likely effects of an Agency action on the ecosystem and how this will result in changes in

the provision of ecological services. The committee recommends that all ecological

valuations conducted by EPA be sufficiently supported by ecological modeling and

ecological data designed to provide insight into or estimates of the likely ecological impacts

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associated with major alternatives being considered by decision makers. The committee

recognizes that EPA is strengthening its approach for developing and using models for

decision-making. For example, EPA has established the Council for Regulatory

Environmental Modeling (CREM), a cross-Agency council of senior managers with the goal

of improving the quality, consistency, and transparency of models used by the Agency for

environmental decision making. The committee endorses this effort and advises EPA to

make effective ecological modeling one of its priorities.

Since many ecological models exist to choose among and for any particular valuation
process a variety of ecological models might be utilized, the Agency will often be faced with
selecting one or more predictive models for use in operationalizing the conceptual model.
The appropriate choice of models, and the availability and appropriateness of supporting
databases, will be different depending on the scale of analysis (e.g., local vs. national) and
the precision of the analysis related to the relevant policy decision. The committee
recommends that EPA identify clear criteria for selection of ecological models for use in
ecological valuation and that the Agency apply these criteria in a consistent and transparent
way.

Several available reports discuss the selection and use of models for environmental
decision making, and the committee believes that these can provide valuable guidance to
EPA regarding criteria for model selection. In 2005 EPA's Council for Regulatory
Environmental Modeling prepared a "Draft Guidance on the Development, Evaluation and
Application of Regulatory Environmental Models." In 2006, an EPA Science Advisory
Board panel reviewed the draft report and provided recommendations on how it should be
revised (U.S. Environmental Protection Agency Science Advisory Board 2006). A final
report is expected. Until the final guidance is published, the original draft guidance and SAB
review can provide the EPA with valuable advice in the selection of models. Similarly, in
2007 the NRC Board on Environmental Studies and Toxicology published a report entitled
"Models in Environmental Regulatory Decision Making." The EPA should utilize this NRC
report as a primary guidance document in selecting appropriate ecological models for use in
valuation exercises. Criteria such as these can guide the Agency both in selecting from
among existing models and in setting priorities for future model development.

The reports discussed above address environmental modeling in general and do not

focus on the use of ecological models for valuation purposes. Thus, in addition to the criteria

discussed in these reports, at least one other criterion specific to the valuation context should

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be considered. The committee recommends that EPA selects predictive models for use in

valuation, the Agency should choose models that generate outputs in terms of the important,

highly valued ecological services identified in the conceptual model or outputs that are easily

translatable into effects on such services. This will greatly facilitate the valuation of

ecological effects. Thus, when using the reports referred to above, the EPA should keep in

mind that the ultimate goal is to provide a measure of the value of the effects of an action on

ecological services.

3.3.2. Identifying Ecosystem Services

Another challenge in implementing ecological production functions in a specific

valuation context is identifying the relevant outputs, i.e., the ecosystem services. The
discussion in the previous section relates primarily to using ecological science to model and
understand the ecology underlying the ecosystems impacted by EPA actions and to predict
ecological changes stemming from those actions. As illustrated in Figure 4, to be useful for
valuation, these changes must ultimately be linked to changes in ecosystem services through
an ecological production function. However, the relevant services must first be identified in
a consistent and appropriate way.

Throughout this report, the committee uses the term "ecosystem services" to refer
broadly to the ecological characteristics, functions, or processes that directly or indirectly
contribute to the well-being of human populations (or have the potential to do so in the
future). This definition includes the intermediate and end products that ecosystems provide.
Regardless of how ecosystem services are defined, the key point is the identification of a set
of changes to ecosystem components that will be valued in a way that is meaningful in the
specific context of interest. For example, if a given ecological change reduces the population
of bees, which in turn reduces pollination, then one would want to value the change in
pollination by comparing or characterizing human well-being with and without the change.
Similarly, if an ecological change increases habitat suitable for a particular species or
activity, one would want to value the change in habitat by comparing human well-being with
and without the change.

Identifying the relevant ecosystem services cannot be done deductively; it is
dependent upon what is important to people, once they have been informed about potential
ecological effects. The ultimate goal is to identify what matters in nature and to express this
intuitively and in terms that can be commonly understood. Technical expressions or

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descriptions meaningful only to experts are not sufficient; similarly, the identification of

relevant services must be informed by the underlying ecological science. Thus, the

identification of relevant services requires a collaborative interaction between ecologists,

social scientists, the public and stakeholders. Input from the public and from stakeholders

can come from a variety of sources, such as the valuation methods described later in this

report (e.g., surveys, individual narratives, mental model research, and focus groups) or from

content analysis of public comments, solicitation of expert opinion and testimony, and

summaries of previous decisions in similar circumstances. The Millennium Ecosystem

Assessment (Millennium Ecosystem Assessment Board. 2003 provides a good starting point

for this exercise by providing an extensive discussion and classification of ecosystem

services.

The committee believes that moving toward defining ecological impacts in terms of
changes in services or ecosystem components that are commonly understood is a key to
success in valuing the protection of ecological systems and services, and urges the Agency to
promote efforts to move in this direction. The relative success of EPA efforts to translate air
quality problems into human health-related social effects is due in part to the development of
agreements about well-defined health outcomes that can then be valued. In order to value the
health effects of air pollution, it was necessary to move from describing impacts in terms
such as oxygen transfer rates in the lung to terms that were more easily understood and
valued by the public, such as asthma attacks. The search for common health outcomes that
can be used for valuation has been difficult. Nevertheless, the lesson is clear: if health and
social scientists are to productively interact (e.g., to assess the economic value of improved
air quality), measures of health outcomes that are understandable and meaningful to both are
necessary. These outcomes are now understood by disciplines as divergent as pulmonary
medicine and urban economics (U.S. Environmental Protection Agency Science Advisory
Board, 2002). The search for common outcomes that can be valued will be especially
important in the ecological realm, where biophysical processes and outcomes can be highly
varied and complex.

Some authors have advocated the development of a common list of services to be

collectively debated, defined and used by both ecologists and social scientists across contexts

(e.g., Boyd and Banzaf, 2007). Such a list might include: species populations (e.g., including

those that generate use value - such as harvested species and pollinator species - and those

that generate existence values); land cover types (e.g., forests, wetlands, natural land covers

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and vistas, beaches, open land and wilderness); resource quantities (e.g., surface water and

groundwater availability); resource quality (e.g., air quality, drinking water quality, soil

quality); and biodiversity. These services play a role in a variety of contributions to human

well-being provided by ecosystems.

Although only a subset of the services on a common list might be relevant in any
particular context, the list would provide some standardization in the definition of ecosystem
services across contexts. Advocates argue that development of a common list is the only way
to debate and convey a shared mind-set, and that it will concretely foster the integration of
biophysical and social approaches and provide greater transparency, legitimacy, and public
communication about what in nature is being gained and lost. While achieving agreement on
a common list might be an important ultimate goal, it is likely to be difficult for complex
ecological systems. Converging prematurely on a limited list of services could misdirect
valuation efforts and miss important intermediate and end services.

The identification of relevant ecosystem services, either as a common list or for a
specific problem, should follow some basic principles to ensure that the services identified
capture socially important ecological changes. These principles include the following:

a)	In identifying the relevant services to be valued, it is important to include all
ecosystem services, but avoid double counting. Here the principle is to count
all things that matter, but to count them only once. The conceptual model
developed to guide the valuation process should be designed to ensure that
this principle is followed.28 In identifying and listing the ecosystem services
to be valued, it is important to capture both intermediate and final services of
importance, recognizing that ecological functions or processes are generally
inputs into the production of another ecological good or service.

b)	Ecological services should have concrete outcomes that can be clearly
expressed in terms that lay publics can understand. In order to provide useful
input into valuation, ecological outcomes must be described in terms that are
meaningful and understandable to those whose values are to be assessed.

Thus, ecosystem services need to be identified through interactions between
technical experts and lay publics. This will involve input from both the
scientific community and from a wide range of interested parties, as a means
of validating the relevance of the services.

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c)	The delineation of services should reflect the basic principles of ecology. In

particular, the delineation should reflect the role of spatial and temporal
phenomena and the importance of place. In practice, the delineation means
that they should be derived from processes that take place at large spatial and
temporal scales, but they should be expressed in local terms at specific times.
For example, the availability of water in a particular place at a particular time
is what people care about, but landscape-level and inter-temporal analyses are
necessary to predict changes in that specific service. Advances in information
technology, mapping, and remote sensing technologies in particular will
increasingly enable this kind of measurement.

d)	The delineation of ecological services should reflect scarcity, and the
availability of substitutes and complements. This is related to the need for
spatially- and temporally-explicit services. The social value of ecological
changes will often be related to the existence of substitutes and complements.
Is this the only clean lake people can swim in or are there others nearby? If
people want to hike in the woods, are there trails they can use? If people like
to kayak in June, will there be adequate water volume? These are often key
determinants of the value of a change. Services should be defined so as to
allow a consideration of scarcity, substitutes, and complements in estimating
or characterizing values.

Figure 4 distinguishes between ecological endpoints and the concept of ecosystem
services, and highlights the fact that identifying ecological endpoints is not the same as
identifying ecosystem services. EPA has several on-going initiatives related to ecological
endpoints, but these fall short of identifying ecosystem services, mainly because they do not
follow the basic principles outlined above.

One ecological endpoint initiative is the Environmental Monitoring and Assessment

Program (EMAP), which the Agency created in the early 1990s. It was designed to be a

long-term program to assess the status and trends in ecological conditions at regional scales

(Hunsaker and Carpenter 1990; Hunsaker 1993; Lear and Chapman 1994). Referring to

EMAP, the EPA recently stated that, "A useful indicator must produce results that are clearly

understood and accepted by scientists, policy makers, and the public." (Jackson et al. 2000)

While this goal is consistent with the goals underlying the identification of ecosystem

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services, the indicators developed in EMAP are not generally direct measures of ecosystem

services. Authors have noted the need to translate EMAP indicators "into common language

for communication with public and decision-making audiences" (Schiller et al. 2001.). In

one analysis, focus groups were used to evaluate the indicators. In general, the study

demonstrates the need "to develop language that simultaneously fit within both scientists'

and nonscientists' different frames of reference, such that resulting indicators were at once

technically accurate and understandable." The committee agrees with this conclusion, and

urges EPA to move toward this goal.

EPA has also developed a set of Generic Ecological Assessment Endpoints (U.S.
Environmental Protection Agency Risk Assessment Forum 2003) based on legislative,
policy, and regulatory mandates. If expanded to include landscape-, regional-, and global-
level endpoints (see U.S. Environmental Protection Agency Risk Assessment Forum 2003
Table 4.1, Harwell, et al. 1999; Young and Sanzone, 2002), the GEAEs can be used as a first
step in characterizing the relevant ecological system and quantifying the responses to
stressors. Thus, the committee views these initiatives as steps in the right direction.

While the GEAEs are a starting point, they also are an example of how far EPA must
go in moving toward consideration of impacts on ecosystem services. First, the GEAEs are
expressed in technical terms and do not generally describe concrete outcomes and are not
expressed in terms that the lay public can understand. While these technical terms are
certainly appropriate for some regulatory purposes, most of the public is not likely to be
familiar with them. Hence, they will have limited use in valuation.

Second, the GEAEs do not necessarily reflect the things in nature that people care
about. Although the endpoints were developed via explicit reference to policy and regulatory
needs (U.S. Environmental Protection Agency Risk Assessment Forum 2003 p.5) they depict
a narrow range of ecological outcomes, confined to organism, population, and community or
ecosystem effects. They do not relate to water availability, aesthetics, or air quality, but
rather to kills, gross anomalies, survival, fecundity and growth, extirpation, abundance,
production, and taxa richness. These effects are clearly relevant to biological assessment.
However, for anglers who care about the abundance of healthy fish in a particular location at
a particular time, the lost value from a single dead or diseased fish depends not on the
number of kills or anomalies but rather on how it affects the abundance of healthy fish in the
landscape.

Finally, the GEAEs do not enable analysis of scarcity and the availability of

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substitutes or complements. This is related to the previous limitation. For example, if

anglers care about fish populations because of their impact on catch rates, then the lost value

from a single dead fish in a single lake will depend (among other things) on the scarcity of

fish and availability of substitutes in the relevant vicinity.

The Agency is aware of these issues. The committee raises them primarily: a) to
highlight the difference between the Agency's current approach to defining relevant
ecological endpoints and the committee's vision of ecosystem services, and b) to encourage
the Agency to move toward identification and development of measures of ecosystem
services that are relevant and directly useful for valuation.

The identification of relevant ecosystem services will require increased interaction
within the Agency between natural and social scientists. The committee urges the Agency to
foster this interaction through a dialogue related to the identification and development of
measures of ecosystem services. One vehicle for increased dialogue is through greater
coordination among the Agency's research programs, especially between the Agency's
extramural research programs in ecological research and in Decision-Making and Valuation
for Environmental Policy. The committee believes that these two programs could and should
be more closely linked. A joint research initiative focused on the development of measures
of ecosystem services will address a critical policy need and provide a way for the Agency to
concretely integrate its ecological and social science expertise.

3.3.3. Mapping Changes in Ecological Inputs to Changes in Ecological Services

Once the underlying ecology is understood and modeled and the relevant ecosystem

services are identified, development of the corresponding ecological production functions
will still require a correlation from the ecological inputs to the ecosystem services that those
inputs produce. Although numerous ecological models exist for modeling ecological
systems, as noted above, most of them fall short of what is needed to fully develop this
relationship. Many of these models have been developed to satisfy research objectives, not
Agency policy or regulatory objectives. In the past, outputs of these models have not
generally been cast in terms of direct concern to people, and thus are not designed as inputs
to valuation techniques. They have typically focused on understanding the dynamics in
ecological systems, such as the effect of abiotic driving variables on production, the
interaction among species, and the rate of carbon sequestration on continental scales. For
example, evapotranspiration rates, rates of carbon turnover, and changes in leaf area are

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important for ecological understanding, but have not been translated into values of direct

human importance. This reflects the fact that the links between outputs of some ecological

models and human uses of ecosystems have only recently been a subject of research.

Certainly, there exist some examples of models with outputs directly related to human

values. These include those that predict fish and game populations or forest productivity.

These examples, however, represent a limited set of ecosystem services.

3.4. Strategies to Provide the Ecological Science to Support Valuation

As noted above, the effect of changes in ecosystem structure and functions on the

provision of ecosystem services should be represented by the relevant ecological production
functions; however, implementation of this ideal faces numerous challenges at this time.
Nonetheless, some promising developments suggest approaches that could be used to move
the Agency toward this goal. These include the use of proxies based on functional groupings
or indicators, and the use of meta-analyses. Proxies represent a form of simplification, while
meta-analysis is based on data aggregation. In addition, opportunities exist for improving the
availability of data for use in parameterizing models of ecological systems and the provision
of ecosystem services. These approaches are described briefly below.

3.4.1. Use of Indicators

As noted above, an ecological production function describes ecological inputs and

outputs (i.e., services), and the relationship between them. When a full characterization of

this relationship is not available, some indication of the direction and possible magnitude of

the changes in the provision of services that would result from an Agency action might still

be obtained using indicators. Indicators are measures of key inputs whose changes are

correlated with changes in ecosystem services. In general, an indicator approach involves

selecting key predictive variables or indicators rather than attempting to measure and value

all the possible significant outputs.

To the extent that the indicators used are grounded in ecological science but

expressed in terms relevant for valuation, they can provide information about how ecological

impacts might affect ecosystem services. If it is known that the indicator is positively or

negatively correlated with a specific ecosystem service, then predicting the change in the

indicator can provide at least a qualitative prediction of the change in the corresponding

ecosystem service. In addition, the use of large, complex ecological models can be difficult

pragmatically, especially because of the quantities of required data and the time to

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implement. As a result, making numerous or rapid evaluations is difficult (Hoagland and Jin

2006) and simplification would be far more practical. Thus, the use of indicators that

simplify and synthesize underlying complexity can have advantages in terms of both

generating and effectively communicating information about ecological effects.

Ecologists and environmental scientists have sought to identify indicators of
ecosystem condition that easily can be linked to specific services. Many ecosystem
indicators have been proposed (U.S. Environmental Protection Agency, 1996; National
Research Council, 2000, U.S. Environmental Protection Agency, 2002b, U.S. Environmental
Protection Agency 2007) and several states have sought to define a relatively small set of
indicators of environmental quality to convey the value of ecological services. Indicator
variables have been established for specific ecosystems such as streams (e.g. Karr, 1993) and
for entire countries (e.g. The H. John Heinz III Center for Science, Economics, and the
Environment 2002). The committee acknowledges EPA's work in developing indicators for
air, water, and land and for ecosystem condition and encourages the Agency to see where
those indicators can be linked to specific services relevant to particular decision contexts
where valuation can be useful.

Figure 5 illustrates possible indicators or metrics at different levels of ecological
organization. One type of indicator is provided by functional groupings. Because of their
inherent complexity, ecological systems cannot be characterized in their entirety, nor can
their responses to stressors be completely measured and predicted by single indicators. For
example, because of the large number of species in most ecosystems, it is rarely possible to
list, characterize, or model all of them when attempting to understand the services they
provide. For this reason, ecologists often aggregate large numbers of species into functional
groupings. All members of one functional group are similar in terms of the role that they
play in the ecosystem. For instance, all deciduous tree species might comprise a single
functional group, as might insect-eating birds, or nitrogen-fixing bacteria. The appeal of this
approach is that within a given functional group there may be many different species that
provide a given function even though one or more of the species of the group may not be
present. Changes in the functional grouping can provide an indication of the likely changes
in the associated services even when a precise estimate of the change in those services is not
possible.

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Figure 5: Indicators of Ecological Properties at Different Levels of Organization

Metrics



Ecosystem services



Clean water, disease
¦ protection,
' recreational
opportunities













Ecosystem
functioning



Capture, processing
| and loss of energy
water and nutrients

















Community structure



Temporal and spatial
• distribution and
' abundance of
organisms

Ecosystem
Service
Inputs













Functional groups



The types of functional
roles present, e.g.

producers,
decomposers, etc

















Biodiversity



The kinds and
| numbers of











organisms

Another approach to indicators is designed to incorporate multiple dimensions into a
coherent presentation that describes the status of ecosystems within a region, especially as
they relate to social values and ecosystem services. For example, the "ecosystem report
card" in South Florida (Harwell, et al., 1999) is an example of an indicator based on

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particularly germane criteria, namely, that it: a) be understandable to multiple audiences; b)

address differences in ecosystem responses across time; c) show the status of the ecosystem;

d) characterize the selected endpoints, and e) transparently provide the scientific basis for the

assigned grades on the report card. Through application of these criteria, the indicator is

intended to provide information about the status and trends associated with the ecological

services provided by the South Florida ecosystem. The report card identifies seven essential

ecosystem characteristics that are thought to be important, i.e., habitat quality, integrity of the

biotic community, ecological processes, water quality, hydrological system, disturbance

regime (changes from natural variability), and sediment/soil quality, which were then related

to the goals and objectives for the ecosystem integrity report card.29 Related ecological

outputs were selected based on both scientific issues and societal values. The outputs are not

designed to be monetized, but rather are described by narratives or quantitative/qualitative

grades that are scientifically credible and easily understood by the public. There are other

examples of using report cards to characterize the status of a given ecosystem. The extension

of this idea, of course, is to use changes in the grades as indicators of ecological effects of

EPA actions. The report card approach is a possible method for characterizing contributions

to human well-being for the purposes of Circular A-4 when economic benefits or ecological

services cannot be readily monetized.

Functional groupings provide an examples of possible indicators. Many others exist.
There is currently no agreement on a common set of indicators that can be consistently
applied and serves the needs of decision makers and researchers in all contexts (Carpenter et
al., 2006). However, there are guidelines for specific issues. For example, in evaluating the
economic consequences of species invasion, Leung, et al. (2005) have developed a
framework for rapid assessments based on indicators to guide in prevention and control,
simplifying the ecological complexity to a relatively small number of easily estimated
parameters. Because of the complexity of the interactions between economic and ecological
systems, economists frequently take a similar simplification approach that focuses on effects
occurring only in the relevant markets, assuming that the effects on the broader market are
negligible and can be ignored (Settle et al. 2002).

This simplification approach to ecological modeling will never satisfy those who will

always want to identify all the possible consequences of EPA actions. For example,

Barbier's (2001) study of the economics of species invasion involved a predator-prey model

with inter-specific competition and dispersion. The model results demonstrated that the

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types of ecological interaction determined the extent to which the introduction and spread of

invasive species reduced commercial fishing. He further argues that future models should

consider more complex ecological interactions, habitat modification and non-market

damages (Hoagland and Jin 2006). [Is the suggestion here that Barbier wants to identify all

possible consequences? And is Hoagland and Jin the right reference? Is Barbier arguing this

in a paper by Hoagland and Jin?? KS]

The challenge is the practicality of building ever more complex models that must
address a wide array of issues over multiple spatial and temporal scales. It may well be that
with accumulated experience, it will be shown to be more practical to adopt the simplified
approach of selecting a few key indicators or ecological processes that are correlated with
specific ecosystem services and can be valued. The committee advises EPA to continue
research to develop key indicators for use in ecological valuation. This is likely to be
particularly fruitful when those indicators can be used for key repeated rulemakings or other
repeated decision contexts. Such indicators should meet ecological science and social
science criteria for effectively simplifying and synthesizing underlying complexity while still
providing scientifically-based information about key ecosystem services. In addition, use of
the chosen indicators should be accompanied by an effective monitoring and reporting
program.

3.4.2. Use of Meta-analysis.

A second promising approach to providing information about changes in ecosystem

services is the use of meta-analysis. Meta-analysis or data-aggregation involves collecting
data from multiple sources and attempting to draw out consistent patterns and relationships
from those data about the links between ecological functions or structures and the associated
services. For example, Worm et al. (2006) attempted to measure the impacts of biodiversity
loss on ecosystem services across the global oceans. They combined available data from
multiple sources, ranging from small-scale experiments to global fisheries. In these analyses,
it is impossible to separate correlation and causation, which is a severe limitation. But
examining data from site-specific studies, coastal regional analyses and global catch
databases will allow researchers to draw correlative relationships between biodiversity and
decreases in commercial fish populations—variables that can be monetized.

In a similar data aggregation approach, de Zwart et al. (2006) noted that ecological
methods for measuring the magnitude of biological degradation in aquatic communities are

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well established (e.g. Karr, 1981), but determining probable causes is usually left to a

combination of expert opinion, multivariate statistics and weighing of evidence. As a result,

the results are difficult to interpret and communicate, particularly because mixtures of

potentially toxic compounds are frequently part of these assessments. To address this issue

the authors used a combination of ecological, ecotoxicological and exposure modeling to

provide statistical estimates of probable effects of different natural and anthropogenic

stressors to fish. This approach a) links fish, habitat, and chemistry data collected from

hundreds of sites in Ohio streams; b) assesses the biological condition at each site; c)

attributes impairment [e.g., loss of one or more of 117 fish species] to multiple probable

causes; and d) provides the results of the analyses in simple-to-interpret pie charts. When

data were aggregated from throughout Ohio, 50% of the biological effect was associated with

unknown factors and model error; the remaining 50% was associated with alteration in

stream chemistry and habitat. While the results do not fully explain the biological effect,,

the point is that the technique combines multiple data sets and assessment tools (models) to

arrive as estimates of loss of fish species based on broad patterns. Thus, like the previous

study of the relationship of biodiversity to ocean productivity, this study aggregates data

from many sources and uses various models to arrive at estimates that can be easily

interpreted and at least in the case of game fish species, can be monetized.

3.4.3. Opportunities regarding ecological data

Although data availability is a serious problem in the development of ecological

production functions, data on the structure and function of ecological systems are becoming
more available and better organized across the country. Part of the increased availability is
simply that Web-based publication now enables authors to make data and analysis readily
available to other researchers in electronic forms in electronic format. Also, as governmental
agencies are being held more accountable, data used in decision-making are expected to be
made available to constituents.

Within the ecological research community, the National Science Foundation (NSF)
Long-Term Ecological Research (LTER) program has had an emphasis on organizing and
sharing data in easily accessible electronic datasets. Although these data were rarely
collected for valuing ecological services, they are particularly valuable because they
frequently measure long-term trends. As such, these data are useful in separating short-term
fluctuations from longer-term patterns in ecological properties. Also, the LTER program

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recently has focused on regionalization, in which data are collected from sites surrounding

the primary site, providing a regional context for site-based measurements and models.

Planning for the forthcoming NSF National Ecological Observatory Network (NEON)

includes a Networking Information and Baseline Design (NIBD) component, which connects

the key scientific questions to the data required to answer the questions. The committee

recommends that EPA effectively link into the NEON planning process, and expand its

involvement with the NSF LTER program, which is undergoing a major refreshing of its

research and data sharing protocols.

1.4.4 Transferring Ecological Information

Despite the increasing availability and organization of ecological data, there is rarely
enough available information to support many of these analyses. In addition, the costs are
too prohibitive to allow extensive data to be collected from all the sites on which EPA is
considering action. From an ecological perspective, therefore, an issue arises regarding the
reliability of transferring ecological information, whether from one site to another, or over
different spatial or temporal scales. Information in this sense can include tools or
approaches, data on properties of an ecosystem or its components, and services or
contributions to human well-being provided by an ecosystem.

There are no hard and fast rules for when ecological information can be transferred;
the confidence in doing so depends on the type of information and the system in question.
Given the complexity, the richness of interactions, and the propensity for non-linearity,
extrapolation of ecological information requires caution. Certain generalizations, however,
are possible. Information is more likely to be transferable with greater similarity among
ecosystem contexts. Also, aggregate information, such as data on ecosystem properties, is
more likely to be transferable than information on particular species or the interactions of
particular species. Thus, the ecosystem properties (e.g., leaf area index, primary
productivity, nitrogen cycling patterns) of an oak-hickory deciduous forest in Tennessee
might be transferable to oak-hickory forests in other parts of the eastern United States that
are at similar stages of development. To a lesser extent, the information might be
transferable to other types of deciduous forests.

Information could be transferable to other spatial or temporal scales if the dynamics

over time and space scales are known for the ecosystems. For instance, if data are available

on how the characteristics of an oak-hickory forest change as it develops or goes through

cycles of disturbance, then data transfers from one point in time to another should be

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possible. Similarly, if information is available on how the properties of the system vary with

spatial environmental variation (local climate, soil type, land-use history), then the extension

of information from one spatial context to another should be possible. EPA and other

national and international agencies have sponsored extensive research on the scaling up of

data from particular sites to regions (Citations?). The results from these analyses are

applicable to the transfer of information on ecological properties and services.

To some extent, the same generalizations apply to transferring tools such as models,

although success depends on how generally applicable the tool is and how difficult (in terms

of data requirements) it is to parameterize for other situations. For example, forest ecosystem

models can often be transferred to other forests using available information from sources

such as LTER sites.

3.5.	Directions for Ecological Research to Support Valuation

The committee is aware that EPA plans to redesign a major part of its intramural and

extramural research program to forecast, quantify, and map production of ecosystem services
(see briefings to the C-VPESS, EPA Science Advisory Board 2006c and 2007b)]. Based on
these preliminary briefings, the committee welcomes these efforts as a way to strengthen the
foundation for ecological valuation, although the committee notes with concern the EPA's
limited and shrinking resources for ecological research (EPA Science Advisory Board 2007).
Although the committee has not received any details about Agency plans, it cautions the
Agency to design the research program in a focused way because the cost of implementing
an ecological production function approach in multiple places on multiple issues may be
significant. The committee commends EPA for asking for additional science advice on its
Ecological Research Program Strategy and Multi-year Plan and believes this advisory
activity should be a priority for an SAB panel of interdisciplinary experts in ecological
valuation, drawing on information in the C-VPESS report..

3.6.	Conclusions/Recommendations

Implementation of the C-VPESS valuation process requires prediction of the

ecological impacts of EPA actions, identification of the relevant ecosystem components and
services to be valued, and linking predicted ecological impacts to changes in those
components and services. This is an essential part of valuation and must be done before the
value of those changes can be assessed.

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With regard to predicting ecological impacts and changes in services, the committee

recommends the following:

•	EPA should begin each valuation with a conceptual model designed to
provide a road map to guide the process. A process for constructing the initial
conceptual model should be formalized, recognizing that as an iterative
process, it responds to the addition of new information and multiple points of
view. The conceptual model and its documentation should clearly describe
the reasons for decisions about the spatial and temporal scales of the target
ecological system, the process used to identify stressors associated with the
proposed EPA action, and the methods to be used in estimating the ecological
effects, always recognizing that the selected effects should relate to the
valuation process. In constructing the conceptual model, participation should
be required from staff throughout the EPA, outside experts from the bio-
physical and social sciences, and members of the public who have a standing
in the results of the outcomes

•	EPA should move toward identification and development of measures of
ecosystem services that are relevant and directly useful for valuation. This
will require increased interaction within the Agency between natural and
social scientists. The identification of services should satisfy the basic
principles outlined above, a) counting all things that matter once and only
once; b) expressing outcomes as services that are commonly understood; c)
incorporating appropriate spatial and temporal considerations; and d)
reflecting the role of relevant substitutes or complements, or both.

•	EPA should seek to use ecological production functions wherever possible to
describe how changes in the ecosystem (resulting from stressors created by
different policies or management decisions) ultimately lead to changes in the
provision of ecosystem services.

•	To operationalize the modeling of the ecological processes that will produce
the ecological services, EPA should use predictive ecological models. There
are many ecological models out there. Building on recent efforts within the
Agency and elsewhere, EPA should develop criteria or guidelines for model

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selection that reflect the specific modeling needs of ecological valuation, and

apply these criteria in a consistent and transparent way.

•	EPA should continue and accelerate research to develop key indicators for use
in ecological valuation for key repeated rulemakings or other repeated
decision contexts. Such indicators should meet ecological science and social
science criteria for effectively simplifying and synthesizing underlying
complexity and be associated with an effective monitoring and reporting
program. The Agency should also support the use of methods such as meta-
analysis that are designed to provide general information about ecological
relationships that can applied in ecological valuation.

•	EPA should actively participate in the major efforts to organize ecological
data (e.g., LTER, NEON) both in terms of providing data and in using the
most applicable data sets in its assessments. EPA should promote efforts to
develop data that can be used to parameterize ecological models for site-
specific analysis and case studies or transferred or scaled to other contexts.

•	EPA should carefully plan and actively pursue investments in ecological
research to generate ecological production functions for valuation, including
research funding investments in STAR research on ecological services and
support for modeling and methods development. In addition, the EPA's
National Center for Environmental Research's programs on evaluating
ecosystem services and valuing ecosystem services should be more closely
linked.

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4 METHODS FOR ASSESSING VALUE

The process for implementing the C-VPESS approach requires the use of an
expanded set of methods for characterizing the value of the predicted ecological effects of
EPA actions. This chapter provides information about methods that the committee examined
for possible use in implementing the integrated and expanded valuation process proposed in
Chapter 2 including methods and approaches for transfer of valuation information.

4.1. An Expanded Set of Methods

As noted above, this section discusses methods that the committee examined for

possible use in implementing the integrated and expanded valuation process proposed in
Chapter 2. This list illustrates the variety of methods available and should not be viewed as
exhaustive.

The methods discussed differ in a number of respects, including the underlying
assumptions, the types of values they seek to characterize, the empirical and analytical
techniques used to apply them, their data needs (inputs) and the metrics they generate
(outputs), the extent to which they involve the public or stakeholders, the degree to which the
method has been developed or utilized, the potential envisioned by the committee for future
use at EPA, and the issues involved in implementing the approach.

While these methods are not easily categorized, the committee has organized the
discussion of methods around groupings based on the premises that underlie the methods. In
each case, the goal is to provide the reader with sufficient information about the methods to
allow a preliminary assessment of the role that various methods could play in implementing
the proposed valuation process (including strengths and possible weaknesses of different
methods) and to direct the interested reader to the relevant scientific literature for further
information. The intent is not to provide an exhaustive treatise on any given method.

Table 5 immediately below provides an introduction to these methods. General
descriptions of the categories of methods follow. The concluding section summarizes the
committee's assessments of methods and its recommendations for EPA. Detailed discussion
of specific methods appear in Appendix B of this report. In addition, Appendix B provides
detailed information about the use of survey methods for ecological valuation.

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Table 2: Introduction to Methods Assessed by the Committee

Method

Form of output/units?

What is method intended to
measure?

Source of Information About Value

Reference to
Discussions
in C-VPESS
Report

Does method measure observed behavior,
verbal or written expressions, or progress
related to previously identified goal?

Who

expresses
value?

BIOPHYSICAL RANKING METHODS

Conservation
Value Method

Map of biodiversity, scarcity,
and/or conservation values across
landscape

Contribution to biodiversity

Measurements related to previously
identified goal of biodiversity

Expert -
ecologist or
conservation
biologist

p. 76
p. 200

Embodied Energy
Analysis

Units of free or available energy
from the sun (plus past solar
energy stored as fossil fuels) per
unit of production

Direct and indirect energy
cost of goods and services

Measurements related to previously
identified goal, reduction in energy
depletion

Expert

p. 76-77
p. 210

Emergy

Units of solar energy used to
produce one Joule of a service or
product

Direct and indirect energy
cost of goods and services

Measurements related to previously
identified goal, reduction in energy
depletion

Expert

p. 77
p. 213

Ecological
Footprint

Area of ecosystems required to
produce resources consumed and
to assimilate waste produced

Biologically productive
land area required (directly
and indirectly) to meet
consumption patterns

Measurements related to previously
identified goal, reducing ecosystem
services consumed per unit of land

Expert

p. 77
p. 212

ECOSYSTEM BENEFIT INDICATORS

Ecosystem
Benefit Indicators

Map of the supply of
ecosystems/services showing
quantities of expressed or
estimated demand for those
ecosystems/services across a
landscape

Quantitative but not
monetary approach to
preference weighting for the
ecological effects of policy
options

Measurements related to demand variables
that can be identified by experts or non-
expert lay publics and supply variables as
identified by experts.

Expert and
selected
non-expert
lay public

p. 77
p. 215

MEASURES OF ATTITUDES, PREFERENCES, AND INTENTIONS

Surveys Including
Questions about
Attitudes,
Preferences, and
Intentions

Attitude scales, preference
rankings, behavioral intentions
toward depicted
enviromnents/conditions

Public concerns, attitudes,
values, beliefs, and
behavioral intentions

Verbal reports, choices, rankings, ratings

sample from
public

p. 78

p. 223-255

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relaied lo pre\ iousK identified uoal'.'

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Report

Conjoint Attitude
Survey Questions

Attitudes, preference rankings
implied from expressed trade-off
preferences

Public concerns, attitudes,
values, beliefs, and
behavioral intentions related
to specific trade-offs

Verbal reports, choices, rankings, ratings

sample from
public

p. "8

p. 223-255

Individual
Narratives

Narrative summaries

Implied knowledge, belief
and attitude structures

Verbal report from lay public

sample from
public

p. 79

p. 223-255

Mental Models

Concepts/categorized 'events' in
conceptual models

Causal beliefs and
inferences

Observed decision making behavior,
verbal reports

any

individual
(expert or
non-expert)

p. 79

p. 223-255

Behavioral

Observations of current or prior

Responses to policies,

Past behavior

sample from

p. 78

Observation/Trace

(trace) use of
ecosystems/services

outcomes, and
consequences, in situ



public

p. 223-255

Interactive

Observations of behavior in

Responses to investigator-

Behavior

sample from

p. 223-255

Environmental
Stimulation

simulated/game environment,
implied preferences

controlled changes in
environmental conditions



public



Systems











economic \ii:riioi)s

Market-Based
Methods

Monetary unit: changes in
consumer and
producer surplus

Well-being of individuals in
society, defined as the
individuals' preferences and

Behavior

participants
in the
market

p. 79-80
p. 256

Travel Cost

Monetary unit: WTP as revealed

their willingness to pay for

Behavior

sample from

p. 80



by responses to
differences in travel cost

gains and compensate for
losses



public

p. 260

Hedonic pricing

Monetary unit: marginal WTP as
revealed by responses to
differences in characteristics and
prices of different units of the
product



Behavior

sample from
public

p. 80
p. 263

Averting
Behavior

Monetary unit: WTP as revealed
by responses to

opportunities to avoid or reduce
damages through purchases of
protective goods, substitutes, etc.



Behavior

sample from
public

p. 80
p. 266

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Survey questions
measuring stated
preferences

Monetary Units: WTP, expressed
purchase intentions or in the case
of Conjoint Economic Surveys,
Monetary Units, WTP implied
from expressed trade-off
preferences



Verbal Reports of WTP or responses to
hypothetical choices.

sample from
public

p. su-sip.
p. 269

(ik()l 1' \\l) I'l 15

i.ic i Ai'ki:ssk>\s()i; v \i.i i:s









Focus Groups

Narrative summaries, frequency
tallies, consensus

Full discovery and
articulation of all the values
that are relevant and
exploration of agreements
and conflicts among
stakeholder constituencies

verbal reports

sample from
public

p. 81

p. 283-284

Referenda and

Historical monetary data on

What the body politic as a

Behavior

Selected

p. 81-82

Initiatives

communities' choices regarding
ecological impacts

collectivity values in terms
of policy outcomes



stakeholders

p. 284

Citizen Valuation
Juries

Qualitative summary of jury
decisions which may include
quantitative or monetary
decisions

How a representative group
views the social civil value
of changes to ecological
s\ siems and services

Verbal reports

Selected
stakeholders

p. 82
p. 296

i)i:cisio\-scii:\ci: \m«>\cin:s

Decision-Science
Approaches

Language to be added here

Language to be added here

Language to be added here

Language to
be added
here

page

numbers to
be added

Ml: 111( )l)S I Sl\(

COST \S \ I'ROXY FOR V \l.1









Replacement Cost
(also called
Avoided Cost)

Monetary Units

Cost of replacing ecosystem
services with human
engineered services as an
estimate of value.

Observed behavior

Experts in
engineering

p. 83-84
p. 324

Tradable Permits

Monetary Units

Incremental willingness to
pay for the reductions in
emissions of specific
pollutants covered by the
permits

Observed behavior

Participants
in the permit
market

p. 327

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Report

Habitat

Equivalency

Analysis

Units of habitat (e.g., equivalent
acres of habitat)

Compensation for loss of
ecological services resulting
from injury to a natural
resource over a specific
interval of time

Measurements related to previously
identified goal (e.g., units of habitat)

Experts in
ecology

p. 83-84
p. 328

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4.1.1. Biophysical Ranking Methods

In some contexts, policy makers or analysts are interested in values based on

quantification of biophysical indicators. Possible indicators include species biodiversity,
biomass production, carbon sequestration, or energy and materials use. Quantification of
ecological changes in biophysical terms allows these changes to be ranked based on
individual or aggregate indicators for use in evaluating policy options. Use of a biophysical
ranking does not explicitly incorporate human preferences. Rather, it reflects either a non-
anthropocentric theory of value (based, for example, on energy flows) or a presumption that
the indicators provide a proxy for human value or social preferences. This latter presumption
is predicated on the belief that the healthy functioning and sustainability of ecosystems is
fundamentally important to the well-being of human societies and all living things, and that
the contributions to human well-being of any change in ecosystems can be assessed in terms
of the calculated effects on overall ecosystems health and sustainability. Opinion is mixed
on whether it is an asset or a drawback that these ranking methods are not tied directly to
human preferences.

The committee evaluated two types of biophysical rankings. The first was a ranking
method based on conservation value. This method develops a spatially-differentiated index
of conservation value across a landscape based on an assessment of rarity, persistence, threat,
and other landscape attributes, reflecting the contribution of these attributes to sustained
ecosystem diversity and integrity. The method provides a scientifically based approach to
assigning conservation values that can used by policy makers or stakeholders to prioritize
land for acquisition, conservation or other uses. Based on GIS technology, the ranking
method has the capability to combine information about a variety of ecosystem
characteristics and services across a given landscape, and to overlay ecological information
with other spatial data. In addition, data layers can be used for multiple policy contexts.
Conservation values have been used in various contexts by federal agencies (e.g., Forest
Service, Fish and Wildlife, National Park Service, and Bureau of Land Management) as well
as by non-governmental organizations (e.g., The Nature Conservancy, NatureServe) and
regional and local planning agencies.

The second group of biophysical methods that the committee evaluated was based on
energy and material flows. Energy and material flow analysis is the quantification of the
flows of energy and materials through complex ecological or economic systems, or both.

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These analyses are based on an application of the first (conservation of mass and energy) and

second (entropy) laws of thermodynamics to ecological-economic systems. Examples

include embodied energy, emergy (the available solar energy used up directly and indirectly

to make a service or product), and ecological footprints. Of these three, embodied energy

and ecological footprints are based on a consistent set of principles recognized by the

committee as potentially useful for EPA, while for emergy, some members of the committee

question whether a consistent set of principles appropriate for valuation are used. Embodied

energy measures the (available) energy cost of goods and services using input-output analysis

or flow accounting methods. Ecological footprint analysis also uses input-output analysis,

but measures costs in land units (rather than energy units) based on the biologically

productive land area (rather than the amount of energy) required to meet various

consumption patterns. While such costs can be used to rank alternatives based, for example,

on an energy theory of value, they will provide a proxy for preference-based values only

under limited conditions.

4.1.2. Ecosystem Benefit Indicators

Ecosystem Benefit Indicators (EBIs) offer a quantitative way to illustrate ecological

contributions to human well-being in a specific setting. They use geo-spatial data to provide
information related to the demand for, supply (or scarcity) of, and complements to particular
ecosystem services across a given landscape based on social and biophysical features that
influence (positively or negatively) the contributions of ecosystem services to human well-
being. Examples of these indicators include the percentage of a watershed in a particular
land use or of a particular land type, the number of users of a service (e.g., water or
recreation) within a given area, and the distance to the nearest vulnerable community.

Ecosystem benefit indicators (EBIs) are quantitative inputs to valuation methods.

They can serve as important inputs to valuation methods as diverse as citizen juries and
econometric benefit transfer analysis, which is a monetary weighting technique. EBIs
provide a way to illustrate ecological contributions to human welfare in a specific setting.
The method can be applied to any ecosystem service where the spatial delivery of services is
related to the social landscape in which the service is enjoyed. Existence values (where
spatial location is irrelevant to both provision and value) are the only ecosystem benefit
category where the method would be inapplicable.

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4.1.3. Measures of Attitudes. Preferences, and Intentions

Social and psychological methods seek to characterize the values that are held,

expressed, and advocated by people. They focus on individuals' judgments of the relative
importance of, acceptance of, or preferences for ecological changes. Individuals making the
judgments may respond on their own behalf or on behalf of others (society at large or
specified subgroups). The basis for their judgments could be changes in individual well-
being, or civic, ethical, or moral obligations relevant to ecosystems and ecosystem services.
That is, people may hold, express, and advocate bioecological values or ethical values that
are unrelated or even counter to their own wants and needs.

Social and psychological methods provide scientific means for determining people's
value-relevant perceptions and judgments about a wide array of objects, events, and
conditions. They typically focus on choices or ratings among sets of alternative policies and
may include comparisons with potentially competing social and economic goals. Social and
psychological methods elicit information about preferences and values primarily through
surveys, focus groups, and individual narratives. Experts in this field recently have been
experimenting with eliciting this information through observations of behavioral responses
by individuals interacting with either actual or computer simulated environments.

Surveys typically involve face-to-face, telephone, or mail interviews with large
representative samples of respondents (see Appendix C for a more detailed description of
survey methods). Survey questions are framed as choices (among two or more options),
rankings, or ratings; responses are self-reported by individuals. Survey questions about social
and psychological constructs may include assessments of attitudes, beliefs, and knowledge,
as well as reports of past behaviors and future behavioral intentions. Variations on survey
methods that may be especially useful in assessments of ecosystems and services values
include perceptual surveys (e.g., assessments based on photographs, computer visualizations,
or multimedia representations of targeted ecosystem attributes) and conjoint surveys (e.g.,
requiring choices among alternatives that systematically combine multiple and potentially
competing attributes). Quantitative analyses of responses are usually interpreted as ordinal
rankings or rough interval-scale relative measures of differences in assessed values for the
alternatives offered. Similarities and differences among segments of the public also can be
identified and articulated. Survey questions about social and psychological constructs may
be especially useful when the values at issue are difficult to express or conceive in monetary
terms, or where monetary expressions are viewed as ethically inappropriate. Surveys to elicit

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value-related information have been used extensively by other federal agencies (see

Appendix C for a representative list).

In contrast to surveys that are based on large representative samples, individual
narrative methods - including mental models analyses, ethnographic, and other relatively
unstructured individual interviews - generally employ small, specially selected samples of
informants and analyze responses qualitatively. Rigorous qualitative analyses can expose
subtle differences in individual beliefs and perspectives and the inferential bases of
participant's value positions, as well as identify opportunities for achieving consensus. The
broad class of studies that fall under the umbrella of individual narrative methods can be
particularly useful in identifying unanticipated value perspectives, positions, and concerns
that might be missed by other value-assessment methods.

4.1.4. Economic Methods

The economic approach to valuation is an anthropocentric approach based on utilitarian

principles. It includes consideration of both instrumental values and intrinsic values, but
only to the extent that preservation based on intrinsic value contributes to an individual's
welfare. Because it is utilitarian-based, it assumes there is the potential for substitutability
between the different sources of value that contribute to welfare. In addition, it assumes that
individual preferences, which determine the degree of substitutability for that person, are
well-formed. Most of EPA's work to date on ecological valuation has been based on the use
of economic methods, and these methods are the focus of EPA's Ecological Benefits
Assessment Strategic Plan.

The concept of value underlying economic valuation methods is based on
substitutability, or, more specifically, on the tradeoffs individuals are willing to make for
ecological improvements or to avoid ecological degradation. By itself, an ecological change
that an individual values will increase that person's utility. The value or economic benefit of
that change is defined to be the amount of another good (typically money) that the individual
is willing to give up to enjoy that change (willingness to pay) or the amount of compensation
(typically in money) that a person would accept in lieu of receiving that change (willingness
to accept). The economic benefits captured by this concept of value can be derived not only
from good and services for which there are markets but also from non-market goods and
services. In addition, both use and non-use (e.g., existence) values are included. Thus,
economic valuation captures values that extend well-beyond commercial or market values.

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However, it does not capture non-anthropocentric values (e.g., biocentric values) and values

based on the deontological concept of intrinsic rights. In addition, both willingness-to-pay

and willingness-to-accept measures depend on the individual's current income (as well as

market prices), implying that individuals with higher incomes will typically have higher

economic benefits. This is viewed by many as a drawback of this approach to defining value.

There are multiple economic valuation methods that can be used in principle to
estimate willingness to pay. These include methods based on observed behavior (market-
based and revealed preference methods) methods based on information elicited from
responses to survey questions (e.g., stated preference methods). In contrast, in general
measures of willingness to accept can only be obtained using stated preference methods.

Market-based methods seek to use information about market prices (or market
demand) to infer values related to changes in marketed goods and services. For example,
when ecological changes lead to a small change in timber or commercial fishing harvests, the
market price of timber or fish can be used as a measure of willingness to pay for that change.
If the change is large, then the current market price alone is not sufficient to determine value;
rather, the demand for timber or fish at various prices must be used to determine willingness
to pay for the change. In general, market-based methods are limited to valuing
"provisioning" services supplied in well-functioning markets.

Revealed preference methods exploit the relationship between some forms of
individual behavior (e.g., visiting a lake or buying a house) and associated environmental
attributes (e.g., of the lake or the house). For example, travel cost methods (including
applications using random utility models) use information about how much people implicitly
or explicitly pay to visit locations with specific environmental attributes (e.g., specific levels
of ecosystem services) to infer how much they value changes in those attributes. Hedonic
methods use information about how much people pay for houses with specific environmental
attributes (e.g., visibility, proximity to amenities or disamenities) to infer how much they
value changes in those attributes. In contrast, averting behavior methods use observations on
how much people spend to avoid adverse (environmental) effects to infer how much they
value or are willing to pay for the improvements those expenditures yield.

In contrast to revealed preference methods, stated preference methods infer values or

economic benefits in terms of willingness to pay or willingness to accept from responses to

survey questions. In some cases, survey questions directly elicit information about

willingness to pay (or accept), while under some survey designs (e.g., conjoint or contingent

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behavior designs) monetary measures of benefits are not revealed directly. Rather, some

form of quantitative analysis is needed to derive economic benefit measures from responses

to survey questions. Although the use of stated preference methods for environmental

valuation has been controversial, there is considerable evidence that the hypothetical

responses in these surveys provide useful evidence regarding values.

4.1.5. Group Expression of Values and Social/Civic Valuation

Several methods prove useful in eliciting expressions of values from groups. Focus

group methods elicit information about values and preferences from small groups of relevant

stakeholders engaging in group discussion led by a facilitator. Given the small number of

participants, the goal of a focus group is rarely value assessment per se, but rather an

articulation of all of the values that may be relevant. Use of focus groups early in the

decision process can help in identifying ecosystem effects that might be particularly

important to the public. Focus groups may also be used to develop measurement strategies

for value assessment (e.g., to design a survey).

One type of method focuses on public and group expressions of public value, in
contrast with traditional economic valuation methods that attempt to measure and aggregate
the values that individuals place on changes in ecological systems and services based on their
personal preferences. Using this alternative approach, known as social/civic valuation,
researchers measure the values that groups place on changes in such systems and services
explicitly in their role as citizens. This approach measures the monetary value that groups
place on changes in the systems and services. The groups are asked to evaluate how much the
public as a whole should pay for increases in such systems and services (public willingness
to pay) or should accept in compensation for reductions in the systems and services (public
willingness to accept). The value measurement purposefully seeks to assess the full "public
regardedness" value, if any, that the group attaches to any increase in community well-being
attributable to changes in the relevant systems and services.

Social/civic values, like values based on personal preferences, can be measured either
through revealed behavior or through stated valuations. One principal source of revealed
values for changes in ecological systems and services are votes on public referenda and
initiatives involving environmental decisions. Other public decisions, however, also may
provide measures of social/civil values, including official community decisions to accept
compensation for permitting environmental damage, and jury awards in cases involving

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damage to natural resources. Where revealed values are difficult or impossible to obtain,

social/civil values also can be measured by asking "citizen valuation juries" or other

representative groups the value that they, as citizens, place on changes in particular

ecological systems or services.

Analyses of the outcomes of referenda or initiatives (with or without a follow-up
survey) seek to determine, for example, if the majority of the voting population feel that a
given environmental improvement is worth what it will cost the relevant government body,
given a particular means of financing the associated expenditure. Similarly, analyses of
public votes about whether to accept an environmental degradation (e.g., through hosting a
noxious facility) seek to determine if the majority of the voting population in that community
feel that the environmental services that would be lost are worth less than the contributions to
well-being the community would realize in the form of tax revenues, jobs, monetary
compensation, etc. These approaches provide information about the policy preferences of the
median voter and, under certain conditions,can provide information about the mean
valuations of those who participate in the voting process. The logic of using formal public
outcomes to infer how much society values particular outcomes has been used previously to
estimate the public's willingness to pay (in the form of a commitment of public expenditure)
to reduce mortality rates from health and safety risks.

Like initiatives and referenda, citizen valuation juries provide information on
social/civic values, but they measure stated rather than revealed value. They also incorporate
elements of the deliberative valuation process. Essentially, the group is given extensive
information and, after a lengthy discussion, is usually asked to agree on a common value or
make a group decision. To date, citizen juries have typically been asked to develop a ranking
of alternative options for achieving a given goal. A jury could also be asked to generate a
value for how much the public would (or should) be willing to pay for a possible
environmental improvement, or, conversely, how much it should be willing to accept for an
environmental degradation. Experience with the use of citizen juries for ecological valuation
is very limited to date.

4.1.6. Decision Science Methods

Text to be inserted on Decision Science Approaches for valuing changes in attributes not
readily measured in dollar terms (e.g., they might instead be measured in physical terms,
such as number of birds, or using constructed scales, such as a scale for aesthetics).

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Description would show how Decision Science Approaches allow comparisons by providing

ways to weight changes in attributes using such methods as swing weights, even swaps, and

ratio methods.

4.1.7. Methods Using Cost as a Proxy for Value

A fundamental principle in economics is the distinction between economic benefits

and costs. Economic benefits reflect what is gained by increasing the amount of a given
good or service. Costs, on the other hand, reflect what must be given up in order to increase
a given good or service. Nonetheless, several methods using the cost of producing equivalent
substitutes for an ecosystem service have been used as proxies for value of that ecosystem
service. Methods that use cost as a proxy for value include replacement cost, habitat
equivalency analysis (HEA), and valuing pollution reduction by the price of tradable
emissions permits. Cost methods have gained some popularity, especially in estimating the
value of protecting ecosystems for provision of drinking water or habitat, because it is often
easier to collect information on the cost of providing an equivalent substitute than it is to
provide information on economic benefits. But because costs and economic benefits are two
distinct notions, great care needs to be taken in the application of these methods and in the
interpretation of results using these methods.

The cost of producing a good or service can provide information about the value of
that production only under specific and limited conditions. First, there must be multiple
ways to produce an equivalent amount and quality of ecosystem services. If so, then one
could replace the loss of an ecosystem service via some other means. Second, the value of
the ecosystem service must be greater than or equal to the cost of producing the service via
this alternative means. If so, society would be better off paying for their replacement rather
than choosing to forego the ecosystem services.

An example in which these two conditions may be met is the provision of clean
drinking water for a metropolitan area. Protecting an ecosystem that serves as a watershed
and building a filtration plant may be two ways of providing the same quantity and quality of
drinking water to a city, in which case the loss of watershed protection could be replaced
with a filtration plant. Further, the value of providing clean drinking water for a metropolitan
area far exceeds the cost of a filtration plant to provide it. In this case, one could value the
protection of an ecosystem for the purpose of providing clean drinking water as equal to the
cost of building the filtration plant.

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When these two conditions are met, it is valid to use the cost of providing the

ecosystem services via an alternative means as the value of the loss of one means to produce

ecosystem services. It is important to note that this value is not the value of the ecosystem

services themselves, but only the value of losing one means to produce them. It is not valid

to use cost as a proxy for value, even in this limited sense of value, when these conditions are

not met.

The committee urges great caution in the adoption of methods using incremental cost as
a proxy for value. It must be demonstrated that the conditions for valid use are satisfied and
analyses of incremental costs should not be interpreted as incremental benefits unless these
conditions are met.

4.1.8. Summary and Recommendations

The methods described in this section, and in more detail in Appendix B, were

evaluated by the committee to help the Agency move toward valuations that include an
expanded range of important ecological effects and human concerns. The committee
observes and strongly reminds the Agency that no single method, metric, or index of value
can be used to fully reflect important ecological effects and human concerns for decision-
making, because value is such a complex concept.

The committee advises EPA to follow the "Process for Implementing an Expanded
and Integrated Approach to Ecological Valuation" (Figure 2). High-quality valuations will
follow that proposed process for a specific decision context, will involve a conscious choice
about the types of values to be assessed, and will have transparent communication about the
types of methods used and the uncertainties associated with methods used at different parts of
the valuation process.

Different kinds of decision contexts might call for use of different kinds of methods.
In some cases, the environmental values at stake may principally involve ecosystem services
easily understood by the general public. Recreation services might be involved, for example,
and survey methods or travel cost methods might be appropriate methods to choose. In other
cases, the decision context may involve ecosystem services that are more complex or not
commonly understood by the broader population (e.g., nutrient cycling or biodiversity). In
those instances, decision makers may be interested in what experts value or they might
choose to use mediated modeling efforts to bring experts and lay publics together. In
addition, some types of decisions have different legal constraints affecting the type of

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1	valuation output sought (e.g., economic benefit-cost analyses associated with Regulatory

2	Impact Assessments call for the use of methods that generate economic values wherever

3	feasible) and some methods work better at certain geographic scales (e.g., Habitat

4	Equivalency Analyses at a site-specific scale; Conservation Value Methods at a landscape or

5	regional scale). The choice of method should be appropriate to the decision context and the

6	geographic scale of use. Finally, EPA must consider the cost of using a state-of-the-art

7	valuation method in terms of the information gained for decision making, while operating

8	under Agency budget constraints. Table 3 below briefly summarizes the committee's

9	conclusions regarding methods discussed in this report. It provides cross-references to
10	sections of Appendix B that discuss methods in more detail.

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Table 3: Table Summarizing Methods Discussed in this Report



Degree to Which Method has Been Developed
or Utilized

Potential for Future Use by EPA in an Integrated and Expanded Approach for Valuation

Conservation
Value Method

Components of approach used by

•	U.S. Department of Agriculture

•	U.S. Forest Service

•	U.S. Fish and Wildlife Service

•	National Park Service

•	Bureau of Land Management

•	IUCN

•	The Nature Conservancy

•	Nature Serve

•	Use to focus available conservation funds related to conservation goals

•	Use as a prediction of ecological impacts that would then be used as an input in an
economic valuation study

•	Use in combination with other non-monetary value information (for example, from
social-psychological surveys) to characterize preference-based values when
monetization is not possible or desirable

•	Use as a means of quantifying biophysical impacts when they cannot be quantified
(as required by the OMB Circular A-4)

Embodied

Energy

Analysis

• Has been used by some ecologists and
physical scientists to implement an energy
theory of value

•	When costs can be used as a proxy for value, this method provides information about
ecological values as defined by the energy theory of WHAT?

•	Can be used to rank options or assess impacts in biophysical terms based on required
energy inputs

•	Does not provide an alternative means of monetizing economic values based on
WTP

Emergy

• Has only been used by a small circle of
researchers, some at EPA

• Substantial questions remain about the appropriateness and usefulness as a method
for ecological valuation

Ecological
Footprint

• Has been used extensively by ecologists to
compare resource use by different
populations

•	Most useful as an index of the quantity of ecosystem services consumed

•	Can be used to rank options or assess impacts in biophysical terms based on relative
resource use

Ecosystem

Benefit

Indicators

• The method is new and relatively
undeveloped

•	Input to a wide variety of trade-off analyses (for regulatory analyses or performance
measures)

•	Use as part of public processes designed to communicate the implications of a
change or policy across a variety of scales

•	Use as inputs to economic and econometric methods such as economic benefit
transfer, or stated preference models

•	Use to systematize alternative choice scenarios in choice experiments and stated
preference surveys

Surveys

Including

Questions

about

Attitudes,

• Survey questions measuring social-
psychological constructs are the oldest and
most frequently used methods for
determining public beliefs, concerns, and
preferences

•	Can contribute to initial problem formulation by identifying ecological services and
impacts that most concern citizens and/or identified stakeholders, as well as by
uncovering assumptions, beliefs, and values that underlie that concern

•	Can help to determine socially important assessment endpoints

•	Can be used to assess relative public preferences among policy options

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Degree to Which Method has Been Developed
or Utilized

Potential for Future Use by EPA in an Integrated and Expanded Approach for Valuation

Preferences,
and Intentions

• Survey questions have been and continue to
be used effectively by all levels of
government to measure citizen desires
concerns and preferences

•	Quantitative outcomes may be especially useful when the values at issue are difficult
to express or to conceive in monetary terms or where monetary valuations are
viewed as ethically inappropriate

•	Can be used to help inform and involve publics in decision-making where valuation
has been involved

Conjoint
Attitude
Survey
Questions



• May be especially well-suited for gauging public preferences across sets of complex
multi-dimensional alternatives, likely involved in many EPA regulations and actions
for ecosystems/services protection

Individual
Narratives

• Provides qualitative information and
generally no representative sampling but
may have a role in earlier stages of
valuation

• Can make important contributions to improving the design, development and pre-
testing of more formal surveys that can provide reliable and valid quantitative
assessments of public concerns and values

Mental Models

• Research has focused more on enabling and
informing risk reduction, rather than
motivating or understanding preferences
and trade-offs per se

• Appropriate precursor (i.e., formative analysis) to any formal survey or preference
elicitation method, to improve the validity and reliability of the method

Behavioral

Observation/

Trace

• Relatively new and untested

• Might be used to attain quantitative measures of human use levels useful in
conjunction with economic measures or as separate measures to be correlated with
changes in ecological conditions

Interactive
Environmental
Simulation
Systems

• Relatively new and untested

•	Can engage and communicate with public audiences about what outcomes they
prefer and policies required to achieve those outcomes

•	Respondents can learn through experience about how the ecosystem of interest
responds to various policies or policy aspects and can progressively modify their
expressed policy preferences

Market-Based
Methods

•	Are based on well-established economic
principle and econometric practices

•	Have been used for more than 30 years to
evaluate a variety of economic and
enviromnental policies

• Provides estimate of willingness-to-pay measures of value for the economic
valuation of enviromnental policies (benefit-cost analysis) that affect ecosystem
services that support the provision of goods and services bought and sold in markets

Travel Cost

•	Method is based on well-established
economic principles and econometric
practices

•	Has been extensive use of this method in
analyzing the demand for recreation
services and the value of attributes of

• Provides estimate of willingness-to-pay measures of value for the economic
valuation of enviromnental policies (benefit-cost analysis) that affect ecosystem
services that support the provision of recreation services

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Hedonic
pricing

• Has been widely applied to estimate the
values of site specific amenities and
disamenities as reflected in the prices of
houses

• Provides estimate of willingness-to-pay measures of value for the economic
valuation of environmental policies (benefit-cost analysis) that affect ecosystem
services that affect the market prices of houses

Averting
Behavior

• Substantial literature on the theoretical
dimensions of the method but relatively few
convincing studies demonstrating that it
yields valid estimates in practice

• Provides estimate of willingness-to-pay measures of value for the economic
valuation of environmental policies (benefit-cost analysis) that affect ecosystem
services for which there are substitute activities or goods

Survey

Questions

Measuring

Stated

Preferences

• Extensive literature covering principles and
applications to valuing environmental
changes extending over a 40-year period

•	Provides estimate of willingness-to-pay measures of value for the economic
valuation of environmental policies (benefit-cost analysis) that affect any type of
ecosystem service

•	The only set of methods capable of capturing the economic concepts of non-use
value and existence value

Focus Groups

•	Not clear the extent to which focus groups
are systematically used in EPA policy
making

•	The OMB and other guidelines do not
clearly specify the criteria for using focus
groups

•	Can be useful and important for designing and pre-testing more formal surveys

•	May also contribute to the design of more effective communications of Agency
decisions

Referenda and
Initiatives

• Logic has been used primarily in the
literature on health and safety

•	Can provide monetized values—of the community's formal decision and values,
ceilings, or floors of the median voter's valuation

•	With follow-up surveys can provide information on beliefs, assumptions and
motives regarding the ecosystem preservation issues that voters perceive are at stake

•	Any EPA decision context calling for monetized valuation could employ these
variants, either singly or as cross-checks with conventional revealed preference or
stated preference approaches

Citizen

Valuation

Juries

•	Experimental method in the context of
ecological valuation

•	Used primarily to help governments rank
options for achieving particular goals

•	Only a few efforts have been made to date
to use citizen juries to generate monetary or
other estimates of the social/civic value of

•	Potentially useful both to identify socially important assessment endpoints and to
attach a value, monetary or socio-psychological, to changes in the assessment
endpoints

•	Can expand the role that the public plays in valuations of changes in ecological
systems and service

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environmental changes



Decision

Science

Approaches

• Language to be added here

• Language to be added here

Replacement
Cost (also
called Avoided
Cost)

• The method has been used to provide
estimates of the value of protecting
watersheds for the purpose of providing
clean drinking water

• There is great potential for abuse in using replacement costs to estimate the value of
ecosystem services and it should be used with care

Tradable
Permits

• With the development of tradable permits
for non-market environmental goods, it has
been suggested that the price of a tradable
permit is a proxy for the economic value of
provision of environmental quality or
conservation

• There are no conditions under which the cost of permits could be used as a proxy for
economic value

Habitat

Equivalency

Analysis

•	Originally developed in 1992 to quantity
damages associated with contaminated
wetlands and has since been applied to
cover injuries due to chronic contamination,
spills, and vessel groundings in a variety of
habitats

•	Currently used in Natural Resource
Damages Assessment (NRDA) under Oil
Pollution Action (OPA) And CERCLA
(Superfund)

•	Provides a framing for characterizing bio-physical change

•	Could be used ex ante to compare alternative future actions to identify the action
with the least impact and to compare alternative actions to identify which will yield
the most service or equal service in the shortest time frame

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4.2. Value Transfer

4.2.1. Transfer of Economic Benefits

Economists often use information that allows the measurement of economic

benefits for hypothesized changes in the amount, terms of availability, or quality of
resources that can be derived from a previously conducted valuation study to assign
values to policy-induced changes in another context. This process or method is known as
"benefits transfer." As an example, suppose that a hedonic property value study used
data from the sales of residential homes in Chicago (the study site) to estimate the
incremental change in housing prices associated with variations in the air quality
conditions near these homes. Given a variety of theoretical and statistical assumptions,
measures adapted from the estimates of these price equations have been used to estimate
the marginal value of small improvements in air quality in other cities, such as Cleveland,
New York City, or Los Angeles (the policy site)30. The adjustments that are necessary to
use benefit information from a previous study in a new context depend on a number of
factors, including the needs of each proposed policy application, the available
information about the policy site, and the added assumptions each analyst is prepared to
make.

In light of constraints imposed by the time and money needed to generate original
value estimates, EPA relies heavily on benefits transfer. In fact, benefits transfer is the
primary method EPA uses to develop the measures of economic trade-offs used in its
policy evaluations. Most RIAs and policy evaluations rely on adaptation of information
from the existing literature. EPA's Economic and Benefits Analysis for the Final Section
316(b) Phase III Existing Facilities Rule June 1, 2006 (U.S. Environmental Protection
Agency 2006), EPA's Final Report to Congress on Benefits and Costs of the Clean Air
Act, 1990 to 2010. (U.S. Environmental Protection Agency 1999), and the economic
benefit-cost analysis of the CAFO regulations offer recent examples of policy evaluations
that used benefits transfer methods. While benefits transfer has been used extensively by
EPA for economic values, parallel approaches can and have been used to transfer other
information relevant to ecological valuation (such as information about biophysical
relationships). This section focuses on issues related to economic benefit transfer, but the

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committee notes that similar issues are relevant to the transfer of other types of
information from one application or site to another.

EPA's heavy reliance on benefits transfer raises a significant issue regarding its
validity. Under what conditions can the findings derived from existing studies be
extended to new applications? Inappropriate benefits transfer often is a weak link in
valuation studies. Prior to 2000, the challenges and limitations of benefits transfer
received little attention. This relative lack of attention is surprising, given the prevalence
of benefits transfer in practical valuation efforts, particularly at EPA. Since 2000,
however, a number of environmental economists and other policy analysts have devoted
considerable attention to the issue of benefits transfer, including an entire 2002 special
issue of the journal Ecological Economics (the Wilson and Hoehn [2006] editorial
provides a good overview).

The evaluations of benefits transfer in the literature are uniformly negative. For
example, Brouwer (2000) concludes that "no study has yet been able to show under
which conditions environmental value transfer is valid" (p. 140). Similarly, Muthke and
Holm-Mueller (2004) urge analysts to "forego the international benefit transfer" and
"national benefit transfer seems to be possible if margins of error around 50% are deemed
to be acceptable" (p. 334). However, these evaluations do not do justice to the potential
for careful economic benefits transfer, since they typically adopt a mechanical process to
mimic the steps in an economic benefits transfer. Because benefits transfer is a wide
collection of methods that arise from the specific needs of each policy application, broad
conclusions regarding validity are not meaningful. Rather, assessment of the validity of
the approach requires case-by-case evaluation of the assumptions used in the specific
application of interest, which must consider the similarities and dissimilarities between
the study site and the policy site(s). By this criterion, some applications of benefit
transfer are valid while others are not. For this reason, overall the committee believes
that general conclusions regarding the validity of the application of these methods are not
possible.

4.2.2. Transfer Methods

As noted above, benefits transfer refers to a collection of methods rather than a

single approach. For example, values derived from one or more study sites can be

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transferred to a policy site in three alternative ways. The first is the transfer of a "unit
value." A unit value transfer usually interprets an estimate of the trade-off people make
for a change in environmental services as locally constant for each unit of change in the
environmental service. For the policy site the relevant (and available) values for these
factors would be used to estimate an adjusted measure for the unit value based on the
specific conditions in the policy area (see Brouwer and Bateman 2005 for another
example in the health context). As noted above, the required adjustments will depend on
a number of factors (see further discussion below).

The second approach is the "function transfer" approach, which replaces the unit
value with a summary function describing the results of a single study or a set of studies.
For example, a primary analysis of the value of air quality improvements might be based
on a contingent valuation survey of individuals' willingness to pay to avoid specific
episodes of ill health (i.e. a minor symptom day such as a day with mildly red watering
itchy eyes; a runny nose with sneezing spells; or a work-loss day described as one day of
persistent nausea and headache with occasional vomiting).31 A value function in this
context would relate the willingness to pay to respondent characteristics and other factors
that are likely to influence it, such as income, health status, demographic attributes, and
the availability of health insurance. This value function could then be used to estimate
willingness to pay for populations with different characteristics. Alternatively, the
original study might estimate a demand function or discrete choice model based on an
underlying random utility model describing revealed preference choices. The demand
function or discrete choice model would be transferred and then used to estimate
economic benefits at the policy site. In this case, the function being transferred would be
an estimated behavioral model rather than a value function. Meta-analyses, which
statistically combine results from numerous studies, also involve a type of function
transfer. Meta-analyses can be undertaken when there is accumulated evidence on
measures of economic tradeoffs for a common set of changes in resources or amenities.
One area with a large number of applications is water quality relevant to recreation [see
Johnston et al. (2003) as an example of meta-analyses for water quality; Smith and Kaoru
(1990a, 1990b) for other recreation-based meta-analyses]. This approach was used
recently in EPA's assessment for the Phase III component of the 316(b) rules.32

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The third approach to benefits transfer is the "preference calibration" approach.
It uses information from the study site to identify the parameters that describe underlying
preferences, with the objective of then using the resulting preference relationship to
estimate benefits at the policy site. With calibration, not all relevant parameters (in this
case relating to preferences) are estimated directly from the data. Rather, some are
calculated or inferred from available estimates of other parameters and assumed or
observed relationships and constraints. With a specified algebraic function describing a
preference relationship, along with information about the factors constraining an
individual's choice in the study application and in the policy application, the preference
calibration approach considers whether there is sufficient information in existing
estimates to calculate or infer the relevant parameters of the preference relationship.

When the parameters can be calibrated or estimated from the existing literature, the
transfer involves using the calibrated preference function, together with the conditions at
the policy site, to measure the trade-off for the change associated with the policy
application The task does not require that the parameters required for all possible trade-
offs (i.e., the complete preference relationship) be calculated, but only those parameters
that are needed to construct a set of trade-offs associated with the economic benefit
measures that are necessary for the policy analysis.33 This technique imposes specific
requirements on the information from existing studies. As a rule, these information needs
are defined by the trade-off concepts measured in the literature (see Smith et al. [2002]
for an example).

4.2.3. Guidance Regarding Benefits Transfer

Regardless of the type of transfer method used, in general economic benefits or

economic value functions derived from a particular ecosystem study site should not
necessarily be expected to be relevant for a particular policy site. Differences in both
biophysical characteristics and human values dictate that great care must be taken in
deciding whether the valuation of economic benefits in one context can be validly used in
another context.

The challenge of transferring benefit estimates is exacerbated by the fact that
often few economic benefit studies are available for a given ecosystem, thereby limiting
the set of comparable cases. One consequence is that analysts sometimes rely on

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1	estimates that are too old to be reliable for new applications. For example, the RIA

2	conducted for the CAFO rule based its willingness-to-pay estimates for improved water

3	quality on indices taken from a contingent valuation study (Mitchell-Carson) that was

4	more than 20 years old. In addition, due to lack of suitable previous studies, analysts

5	might inappropriately use values or functions derived from studies designed for purposes

6	other than those of the policy site. For example, the Mitchell-Carson study used in the

7	CAFO RIA was not intended to apply to specific rivers or lakes. Moreover, the water

8	quality index used by Mitchell and Carson was highly simplified, with no intention of

9	capturing the ecosystem services beyond those related to fishing.

10	An additional challenge stems from the difficulty of finding the most appropriate

11	unit values to carry over from the study site to the policy site. As the example in Text

12	Box 1 shows, several different metrics of value are possible (e.g., the number of fish

13	anglers catch per outing; the number of fish caught per hour), and the different metrics

14	will have very different implications for the valuation at the policy site. The choice of

15	unit values has to be appropriate to the scale and context as well. For example, the

16	willingness to pay for increased wilderness areas in a study site may have been expressed

17	in terms of dollars per absolute increase in area (e.g., $100 per taxpayer annually for a

18	100-acre increase in area, or $1 per acre). This unit value may be reasonable for a small,

19	heavily populated municipality, but far too high for a municipality with much more

20	existing wilderness area.

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37

Text Box 1: The Challenge of Choosing a Unit Value for Economic Benefits Transfer

Suppose estimates from the literature imply that the average value of the
willingness to pay for a 10% improvement in the catch rate (i.e. fish caught per
unit of effort) for a sport fishing trip is $5 per trip. This estimate could be from
one study describing specific types of fishing trips by a sample of individuals or it
could be an average of several studies.

One approach for developing a unit value transfer would divide $5 by 10%
to generate a unit value of $0.50 for each 1% improvement. This strategy
implicitly assumes the benefit measure is not influenced by the level of the
quality. It is assumed to be constant for each proportionate improvement. Another
approach would take the same information on average trade-offs and calculate a
unit value using the level of the quality variable -in this case a catch rate which
itself embeds another economic decision variable -the effort a recreational fisher
devotes to fishing. For the example the quality or number of fish caught per hour
of effort must be known. Suppose that in the study providing the estimated
economic benefit the average number of fish caught with an hour of effort before
the improvement was 2. Thus a 10% improvement means that the typical

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recreationist would catch 0.2 more fish with an hour's effort, implying a unit
value of $5 for every additional 0.2 fish caught per hour of effort, or (assuming a
linear relationship in terms of the catch rate rather than the proportionate change
in this quality measure) $25 for every additional fish caught per hour of effort.

Finally, the unit value could be expressed in terms of improved fishing
trips. Suppose the average recreational trip involves 5 hours of fishing over the
course of a day. Then the improvement of 0.2 fish per hour implies an average of
one more fish is caught during a trip. These additional data of the features of the
trips might be used to imply that the improvement made typical trips yield
incremental economic benefits of $5 per trip (the value of catching 0.2 additional
fish per hour for a period of five hours). There are other ways this estimate could
be interpreted. These examples are not intended to be the only "correct" ones or
the best. They illustrate that the information on the baseline conditions, the
measurement of quality and the measurement and terms of use all can affect how
a given set of estimates is used in a benefits transfer.

For the study site all three interpretations are simply arithmetic
transformations of the data describing the context for the choices that yield the
trade-off estimates. However, the same conclusions do not hold when they are
transferred to a different situation. Suppose the policy site involves a case where
we wish to evaluate the effects of reducing the entrainment of fish in power plant
cooling towers. Assume further it was known from technical analysis that this
regulation would lead to 5% improvement in fishing success along rivers affected
by a rule reducing fish entrainment. If these areas have 2,000 fishers, each taking
about 3 trips per season and currently they catch 1 fish per hour, the alternative
unit value transfers would be:

Table 4: Table of Alternative Unit Value Transfers

Assumption

Unit Value

Interpretation of
Policy

Aggregate Value

Constant Unit
value for a 1%
improvement

$0.50 per
1%

improvement

5% improvement
per trip

$0.50 * 5 * 3 * 2000 =
$15,000

Constant unit
value for an extra
fish caught per
hour of effort

$25 per
additional
fish per hour

Added fish caught

$25 * .05 * 1 * 3 * 2000
= $7,500



Constant Value
"or an "improved"
trip

$5.00 per
trip

improved fishing
trips

$5 * 3 * 2000 = $30,000



Clearly these examples deliberately leave out some important information. Trips
may be different - longer, require more travel time, or involve different features
such as different species or related activities. These added features were aspects
that were omitted in the example. These estimates also do not allow for the
possibility that fishing success induces existing recreationists to take more trips
and or that people who never took trips may start taking them after the
improvement. Under each of these possible outcomes, the sources for error in the

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transfer compound Iacii uillioul such details, these simple examples illustrate

how the aggregate economic benefit measures differ by a factor of four.

Two approaches can address the challenges of determining whether and how a
benefits transfer should be conducted.: a) developing guidance for the analyst to
determine whether a value derived from a previous analysis ought to be transferred; and
b) creating procedures to ensure the appropriateness of the choice of study site(s), the
assumptions underlying the methods used, and the resulting values.

The broad categories for evaluating the appropriateness of economic benefits
transfer arise from understanding that how people value the preservation or alteration of
an ecosystem depends on two dimensions: (a) their preferences and (b) the nature of the
biophysical system. The similarities or differences expected in preferences are likely to
depend on how close the stakeholders in the two cases are along social and economic
dimensions that influence the marginal willingness to pay (MWTP). For example,
income levels or age profiles are sometimes relevant, as in many cases of valuing
recreational opportunities. The particular cultural characteristics of the community also
may be relevant. For example, in locations where salmon are seen as iconic species
reflecting the entire ecosystem (e.g., Seattle), people are likely to value salmon more
highly, and are more likely to value the water quality attributes regarded as important for
preserving the salmon stock.

When only information on willingness to pay per unit of improvement is
available, the analyst must be sensitive to the types of differences that would render the
transfer inappropriate. If all the differences between the study site and the policy site are
such that one is likely to have a higher value per unit of improvement than the other, the
study site can effectively provide either a floor or ceiling for the policy site. When the
information from the study site is in functional terms (e.g., willingness to pay as a
function of income levels or age), socio-economic differences between the study site and
the policy site can be accommodated, if these specifications are valid.

Although it may be possible to adjust for differences in socio-economic
characteristics of the populations, major biophysical differences will affect the value even
if every individual in the study case were matched by one in the policy case (e.g., the

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value of improving the water quality of one small lake in Minnesota compared to Texas).
Therefore, the capacity to adjust for biophysical differences is typically more limited.

4.2.4. Screening Process

This procedural approach is based on the premise that a deliberate effort to

examine the similarities and differences between study sites and the policy site, by both
EPA analysts and those providing oversight of their work, will help to flag problematic
transfers and clarify the assumptions and limitations of the study site results. Several
procedures can be considered, one of which is to contact experts familiar enough with
both the previous and current contexts to determine whether to proceed with the
economic benefits transfer. These experts presumably will apply the criteria that they
regard as relevant, even if the set of criteria are not explicit. Experts knowledgeable in
both the study case and the policy case can suggest the most appropriate functional forms
and unit values. For example, Desvousges, Johnson and Banzhaf (1998) relied on expert
judgments to convert estimates of trade-offs to avoid health-related symptoms into the
implied trade-offs expressed in terms of changes in an index of the quality of life (i.e. the
quality of well-being). Experts may also be able to suggest other existing valuations that
would be better candidates for transfer of willingness-to-pay or willingness-to-accept
information.

Another procedure is to make a detailed examination of the appropriateness of the
study case part of the regular routine of the in-house review of EPA analyses using
benefits transfer. Such oversight would require the analysts to clarify the assumptions,
purposes, and units of the study-site analysis so that the in-house reviewers could judge
the appropriateness of the transfer. Analysts must also be fully transparent regarding the
origin and date of the original valuation.

More thorough cataloguing of existing valuation studies, with careful descriptions
of the characteristics and assumptions of each, would be helpful in increasing the
likelihood that the most comparable existing valuations will be identified. This is an
additional rationale for developing databases of valuation studies. The establishment of a
Web-based platform for data and models focusing on valuation estimates would be very
worthwhile. Comparable to the Web sites developed and maintained for other large scale
social science research surveys such as the Panel Study on Income Dynamics (PSID) and

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the Health and Retirement Study (HRS), such a platform could expand the ability of
Agency analysts to search for the most appropriate study cases and to supplement these
records with related data for transfers. While some limited efforts along these lines are
currently underway (see, for example, the database being developed for recreational use
values — http://www.cof.orst.edu/cof/fr/research/ruvd/Recreation_Letter.html), a
systematic effort across a wide range of ecosystems services is needed.

4.2.5. Recommendations

The committee advises EPA to explicitly identify relevant criteria related to

societal preferences and the nature of the biophysical system of the cases being

considered for economic benefits transfer to determine the appropriateness of the transfer.

Both EPA analysts and those providing oversight of their work must take into account the

differences between study site and policy site to flag problematic transfers and clarify the

assumptions and limitations of the study site results.

The committee also advises EPA to develop a Web-based catalogue of existing

valuation studies across a range of ecosystem services, with careful descriptions of the

characteristics and assumptions of each, to assist in increasing the likelihood that the

most comparable existing valuations will be identified.

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5 CROSS-CUTTING ISSUES

This chapter addresses two topics important to ecological valuation in all decision
contexts contexts: analysis and communication of ecological valuation information. The
sections below describe special issues related to ecological valuation and committee
recommendations about how they can best be addressed by EPA.

5.1. Analysis and Representation of Uncertainties in Ecological Valuation

5.1.1. Introduction

Ecosystem valuation efforts are subject to uncertainty, regardless of the method
used. Assessments of uncertainty allow more informed evaluations of proposed policies
and comparisons among alternative policy options. Because any given policy may result
in a range of different outcomes, decision makers must be provided with sufficient
information regarding what is known about the distribution of possible outcomes so that
uncertainty can be taken into account when they make their policy choices. By
identifying key uncertainties, it is also possible to develop potentially important insights
regarding the design of research strategies, thus reducing uncertainty in future analyses.

When reflecting on the role of uncertainty in ecological valuation, three key
questions arise. First, what are the major sources of uncertainty? More specifically, what
types of uncertainty are likely to arise when using alternative valuation methods for
specific applications? Second, what methods are available to characterize and
communicate uncertainty in the results of ecological valuations? A third and final key
question is associated with the types of research - data collection, improvements in
measurement, theory building, theory validation, and others - that can be pursued to
reduce uncertainty for particular sources in specific applications. Section 2 briefly
describes the major sources of uncertainty in ecosystem and ecosystem services
valuation. (See Appendix B for more specific discussions of the uncertainty arising from
the use of specific valuation methods.) Section 3 then discusses two approaches to
characterizing or communicating uncertainty regarding ecological values, namely, Monte
Carlo analysis and expert elicitation. Finally, Section 4 discusses how uncertainty
analysis can be used to set research priorities.

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5.1.2. Sources of Uncertainty in Ecological Valuations

Valuation of the contributions to human well-being of proposed public policies

entails three analytic steps, each potentially subject to uncertainty: predicting biophysical
outcomes; predicting behavioral reactions to these outcomes; and valuing the
consequences of all of these changes. It might be tempting to limit attention to
uncertainty in the third step where values are ultimately estimated, but the uncertainties in
each stage of the analysis are of potential importance, and there is no reason - on the
basis of theory alone - to judge one to be more important than the other a priori. Rather,
the relative magnitude of the uncertainty involved in each step in the valuation process is
fundamentally an empirical question.

At each stage, uncertainty can arise from several sources. First, some of the
physical processes might be inherently random or stochastic. Second, there can be
uncertainty about which of several alternative models of the process best captures its
essential features.34 Finally, there are uncertainties involved in the statistical estimation
of the parameters of the models used in the analysis.

For example, at the biophysical level, any characterization of current (or past)
ecological conditions will have numerous interrelated uncertainties. Any effort to project
future conditions, with or without some postulated management action, will magnify and
compound these uncertainties. Ecosystems are complex, dynamic over space and time,
and subject to the effects of stochastic events (such as weather disturbances, drought,
insect outbreaks, and fires). In addition, our knowledge of these systems is incomplete
and uncertain. Errors in projections of future states of ecosystems are thus unavoidable,
and constitute a significant and fundamental source of uncertainty in any ecological
valuation.

Every social, economic, or political forecast is also based on implicit or explicit
theory of how the world works, either represented by the "mental models" of the
forecasters or by the "mental models" underlying the formal and explicit methods used in
econometric modeling, systems dynamics modeling, and other forms of modeling.
Theories and their expressions as models are unavoidably incomplete and may simply be
incorrect in their assumptions and specifications.

Valuation methods per se are also subject to data and theory limitations. They
unavoidably rely on assumptions that introduce uncertainty. The uncertainties that arise
with various methods are discussed in Appendix B. In addition, analysts are often

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required to apply estimated values to contexts that differ from those in which they were

developed. The possibility that appropriate adjustments are not made in transferring

estimates to different contexts introduces another source of uncertainty. .

In order to identify the types of uncertainty most likely to be at issue for
individual valuation approaches in specific contexts, two issues are relevant: the
sensitivity of an approach to the potential sources of uncertainty listed above; and the
magnitude of uncertainty thereby generated. The consequence of data limitations can be
assessed by determining the variation in results implied by variations in data.
Vulnerability to theoretical limitations is more difficult to assess, but can be gauged - in
some cases - by comparing predictions based on alternative models.

5.1.3. Approaches to assessing uncertainty

Probabilistic uncertainty analysis, by its very nature, is complex, particularly in

the context of ecological valuation. The simplest and probably most common approach
to representing uncertainties is some form of sensitivity analysis, which typically varies
one parameter or model assumption at a time and calculates point estimates for each of
the different parameter values or assumptions. The results provide a range of estimates
of the "true" value, including lower and upper bounds. No effort is made to assign
probabilities to the calculated values or estimate the shape of the distribution of values
within the range.

While sensitivity analysis may be sufficient for some simple problems, when used
in the context of ecological valuation it is likely to give an incomplete and potentially
misleading picture of the true uncertainty associated with the value estimates. Due to the
number of sources of uncertainty in many ecological valuations, sensitivity analysis is
unlikely to be able to account for the implications of all the sources of uncertainty. In
addition, this approach becomes unwieldy when the outcomes relevant to the value
assessment are themselves composed of multiple interrelated variables. For example, at
the biophysical level it is extremely difficult to calculate the uncertainty in projecting
outcomes from a complex ecological system composed of multiple interacting variables
subject to the influence of external stochastic events.

Because of the limitations of simple sensitivity analysis, other approaches to
characterizing uncertainty have been developed. These include Monte Carlo analysis and
the use of expert elicitation. These approaches, which are below, will generally provide a

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more useful and appropriate characterization of uncertainty in complex contexts such as

ecological valuation.

Monte Carlo analysis is an approach to characterizing uncertainty that allows
simultaneous consideration of multiple sources of uncertainty in complex systems. It
requires that a model be developed to predict the system's outputs from information
about inputs (including parameter values). In addition, the underlying inputs into the
system that are uncertain are assigned probability distributions. A computer algorithm is
used to draw randomly from all of these distributions simultaneously (rather than one at a
time, as in sensitivity analysis) and to predict outputs that would result if the inputs took
these values. By repeating this process many times, the analyst can generate probability
distributions for outputs, conditional on the distributions used for the inputs.

Developments in computer performance and software over the years have
substantially reduced the amount of effort required to conduct calculations for a Monte
Carlo analysis, once input uncertainties have been characterized. Widely available
software allows the execution of Monte Carlo analysis in common spreadsheet programs
on a desktop computer, with minimal additional effort relative to that needed to produce
point estimates. In developing probability distributions for uncertain inputs, uncertainty
from statistical variation can also often be characterized with little additional effort
relative to that needed to develop point estimates. Much of the data necessary for such
characterizations already will have been collected for the development of point estimates
(although characterizing other sources of uncertainty in inputs can require more effort).

Over the years, the use of Monte Carlo analysis has been shown to provide a more
reliable and rich characterization of the implications of uncertainty than simple sensitivity
analysis. In contrast to sensitivity analysis, Monte Carlo analysis provides information
on the likelihood of particular outcomes within a range. Indeed, an understanding of the
likelihood of values within a range is essential to any meaningful interpretation of that
range. Without such an understanding, inappropriate conclusions may be drawn from the
presentation of a range of possible outcomes. For example, when a range of possible
ecological values is provided, some may assume that all values within that range are
equally likely to be the ultimate outcome, even though this is rarely the case. Others may
assume that the distribution of possible values is symmetric. This, too, often may not be
the case.

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Because of its ability to characterize uncertainty in a more meaningful way,

Monte Carlo analysis has become common in a variety of fields, including engineering,

finance, and a number of scientific disciplines. It has also been found to be useful in

certain policy contexts. In particular, EPA recognized as early as 1997 that it can be an

important element of risk assessments (U.S. Environmental Protection Agency 1997).

However, efforts to formally quantify uncertainties rarely have been undertaken in the

context of ecological valuations. More often, uncertainty has been addressed

qualitatively or through sensitivity analysis.

Despite its advantages, it is unlikely that a Monte Carlo analysis will

comprehensively address all sources of uncertainty in the estimation of ecological values.

Thus, the results of such an analysis will likely understate the range of possible outcomes

that could result from a related public policy. Nonetheless, the ranges produced by such

an analysis would still provide more reliable information about the implications of known

uncertainties than simple sensitivity analysis. In turn, these ranges can better inform

judgments by policymakers as to the overall implications of uncertainty for their

decisions. Thus, the committee urges EPA to move toward greater use of Monte Carlo

analysis as a means of characterizing the uncertainties associated with estimating the

value of ecological protection.

A host of "expert elicitation" methods can provide indications of the amount and

nature of uncertainty associated with estimates of specific values or predictions regarding

impacts of a given activity or change. (See, for example, Morgan and Henri on (1990) or

Cleaves (1994).) In its very simplest form, an expert elicitation is a single expert's

assessment of the uncertainty of an estimate, forecast, or valuation, whether it is based on

implicit judgment or a more explicit approach like the Monte Carlo technique. Policy

makers can elicit more information from the expert, such as the assumptions underlying

his or her analysis or the bases for uncertainty, to better understand the reliability of the

expert's input and the nature of the uncertainty.

Although an elicitation can rely on a single expert, the bulk of expert elicitation

methods involve multiple experts, which allows for a comparison of their judgments and

an assessment of any disagreements. If the experts are of equal credibility, such that no

judgment can be discarded in favor of another, the range of disagreement reflects

uncertainty. That is, if top scientists express strong divergences in their estimates,

forecasts, or valuations, the existence of a high level of uncertainty is irrefutable. This

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relationship, however, is asymmetrical, in that narrow disagreement does not necessarily

reflect certainty. In other words, the experts may all be equally wrong, a somewhat

common occurrence given that experts often pay attention to the same information and

operate within the same paradigm for any given issue (Ascher & Overholt, 1984: 86-87).

When experts interact prior to providing their final conclusions (e.g., by exchanging

estimates and adapting them in reaction to what they learn from one another), the errors

due to incompleteness can be reduced. For example, biologists may benefit from the kind

of information that can be provided by atmospheric chemists, and vice versa. Such

interactions, however, run the risk of "groupthink" - the unjustified convergence of

estimates due to psychological or social pressures to come closer to agreement (Janis,

1982).

For many expert elicitation methods, translation into probabilities is difficult.
Simple compilations of estimates (e.g., contemporaneous estimates of species
populations) from different experts are sufficient to result in a table with the range of
estimates. They are unable, however, to convey the degree of uncertainty that each expert
would attribute to his or her estimate. This information can be conveyed, however, when
the compilation of estimates also includes confidence intervals. The committee believes
that expert elicitation should be used to characterize uncertainty when more formal
uncertainty analysis (e.g., using Monte Carlo methods) is not feasible. In addition, the
committee recommends that EPA use expert elicitation to obtain estimates of parameters
and their uncertainty for use in Monte Carlo analysis, if suitable information about the
relevant range for the parameter values is not available based on observation (e.g., field
work or experiments).

5.1.4. Using Uncertainty Assessment to Guide Research Initiatives

Over time, additional research related to data collection, improvements in

measurement, theory building, and theory validation can reduce the uncertainties
associated with ecological valuation. For example, research can improve our
understanding of the relationships governing complex ecological systems and thereby
reduce the uncertainty associated with predicting the biophysical impacts of alternative
policy options. Even stochastic uncertainty can sometimes be addressed by initiating
research that focuses on factors previously treated as exogenous to the theories and
models. For example, an earthquake-risk model based on historical frequency will have

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considerable random variation due to the exclusion of detailed analysis of fault-line

dynamics; bringing fault-line behavior into the analysis may lead to reductions in such

uncertainty (Budnitz et al. 1997).

In addition, assessments of the magnitude and sources of uncertainty can help to
establish research priorities and to inform judgments about whether policy changes
should be delayed until research reduces the degree of uncertainty associated with
possible changes. Determining whether the major source of uncertainty comes from
weak data, weak theory, randomness, or inadequate methods can help to guide the
decision on how to allocate scarce resources for research, or whether further research is
worth pursuing. Some data needs are simply too expensive to fulfill, and some methods
have intrinsic limitations that no amount of refinement will fully overcome. Uncertainty
analysis can provide insight into whether near-term progress in reducing uncertainty is
likely, based on the sources of uncertainty and the feasibility of addressing these
limitations promptly. However, it is important to avoid the pitfall of delaying a necessary
action simply because some uncertainty remains, since it always will.

5.2. Communication of Ecological Valuation Information

Nearly all of this report focuses on a new conceptual approach to ecological

valuation and the methods and processes for implementing it. Much of the success of the
multi-disciplinary approach described in Chapter 2, however, depends on how EPA
communicates ecological valuation information as it conducts its valuation process.
Although the committee has not devoted extended discussions to the particular
communication challenges presented by ecological valuation, it believes that generally
accepted practices for communication of technical information apply. The committee
also makes several recommendations to help EPA address some of the special
communication challenges that arise for ecological valuation.

Three essential functions of communication in the context of valuing the
protection of ecological systems and services are: a) communication among and between
technical experts and publics within the valuation process itself; b) analysts'
communication of valuation analyses to decision makers; and c) EPA's communication
of the results of the valuation and decision making processes to interested and affected
publics. While at first glance, these communication functions may appear to be separate
steps, they overlap. Success of the overall valuation process and any communication step

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within it depends on understanding how decision makers use valuation information.

Communication spokesmen must understand how different publics and experts have

framed valuation issues before they can communicate the results of a formal valuation

analysis effectively.

5.2.1. Applying General Communication Principles to Ecological Valuation

Any effective communication strategy requires interactive deliberation and

iteration (National Research Council 1996). Effective communication of valuation
information on implementing the conceptual approach to valuation described in this
report where technical experts and interested and affected publics interact to clarify the
values to be represented in the analysis. The potential pool of interested parties for
ecological values include interested and affected publics and scientists, especially
economists, social scientists, and environmental policy scientists. There is likely a broad
public audience interested in better understanding the value of protecting ecological
systems and services, but also an intermediate group of those who would use data and
models, who through their analyses and activities serve as important mediators for this
kind of information. They will need to access technical details and models, as well as
resulting value estimates.

Effective values communication requires systematically supporting interactions
with interested parties, the character of which will differ depending on the technical
expertise and focus of the interested parties. In general, interactive processes are critical
for improving understanding, although messages or reports (such as EPA's Draft Report
on the Environment) are also important, especially in the context of assessment. The
committee recommends that EPA develop an empirical analysis of the users of valuation
and adapt valuation communications to their needs.

Fundamental guidelines for risk and technical communication are generally
applicable to communicating ecological values. To support decisions effectively,
communications must be designed to address the recipient's goals and prior knowledge
and beliefs, taking into account the effects of context and presentation (Morgan et al.
2002). For example, linear graphs are likely to convey trends more effectively than
tables of numbers (Shah and Miyake 2005) and text that incorporates headers and other
reader-friendly attributes will be more effective than text that doesn't (Shriver, 1989).

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Two examples of risk and technical communication guidelines are the

communication principles from EPA's Risk Characterization Handbook (U.S. EPA

Science Policy Committee 2000) and Guidelines for effective Web sites (Spyridakis

2000). The Risk Characterization Handbook principles include transparency, clarity,

consistency, and reasonableness. Spyridakis provides guidance in five categories:

content, organization, style, credibility, and communicating with international audiences.

She provides a concise table worth consulting for those providing information via

websites and provides generally accepted guidance useful for communication of

valuation information, such as:

•	select content that takes into account the reader's prior knowledge.

•	group information in such a way that it facilitates storing that information in
memory hierarchically.

•	state ideas concisely.

•	cite sources appropriately and keep information up to date.

As in the case of any type of communications, it is difficult to predict effects of
communication efforts. Good communications practice requires formative evaluation of
communications as part of the design process. Summative evaluation after the fact will
enable assessments of effectiveness, ultimately leading to continued improvement in
communications (e.g. Scriven, 1967; Rossi et al., 2003). The committee recommends
that EPA evaluate ecological valuation communications to assess their effects and to
learn how to improve upon them.

5.2.2. Special Communication Challenges Related to Ecological Valuation

Although application of these general communication principles will improve

communications relating to ecological valuation, special challenges arise in this context.

As discussed in this report ecological values can be defined qualitatively or quantitatively

and they can be communicated in a wide variety of ways. Several critical design choices

can influence the communication of: a) the ecological functions, systems, and services to

which the valuation pertains; b) the values analyzed - whether to use a quantitative or

qualitative representation and how to accommodate multiple metrics; and c) how to

communicate uncertainty.

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Communicating the value of protecting ecological systems and services requires

conveying not only value information (in terms of such metrics as monetized values,

rating scales, or the results of decision-aiding processes, for example) but also

information about the nature, status, and changes to the ecological systems and services

to which the value information applies. The EPA Science Advisory Board review of

EPA's Draft Report on the Environment (U.S. EPA Science Advisory Board 2005) and

other reports (e.g., Schiller et al., 2001; Carpenter et al., 1999; Janssen and Carpenter,

1999) emphasize that people need to understand the underlying causal processes in order

to understand how ecological changes affect the things they value (e.g., ecological

services).

The latter can be, and often is, conveyed using such visual tools as mapped
ecological information, photographs, graphs, and tables of ecological indicators. To the
extent that such visual outputs, especially outputs from integrated Geographic
Information Systems, using best cartographic principles and practices (Brenner 1993) can
be made interactive, they will facilitate sensitivity analysis that can address audience
questions about scale and aggregation and may be more effective as communication
tools. The U.S. EPA Science Advisory Board has proposed this kind of framework for
reporting on the condition of ecological resources (U.S. EPA Science Advisory Board,
2004). EPA's Draft Report on the Environment (EPA, 2002) and Regional
Environmental Monitoring and Assessment Program reports illustrate a range of
representational approaches.35

The communication of ecological values is complicated by the many uses and
definitions of the term value. The broad usage of the term in this report includes all the
concepts of value described in Table 1 of this report (A Classification of Concepts of
Value as Applied to Ecological Systems and Their Services). A corollary is that people
communicate - and elicit - different kinds values in very different ways, as discussed
earlier in this report. In addition, context and framing can influence strongly how people
rank, rate, and estimate values (Hitlin and Piliavin, 2004; Horowitz and McConnell,
2002), as well as how they interpret value-related information (e.g. Lichtenstein and
Slovic 2006).

As discussed elsewhere in this report, value measures are required or useful in a

variety of regulatory and non-regulatory policy contexts, ranging from national

rulemakings to site-specific decision making and prioritization of environmental actions

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and educational outreach in the context of regional partnerships. In some cases

monetization is required, whereas in others (e.g., educational outreach by regional

partnerships), narratives and visual representations of values appear to play an important

role. Little direct evidence exists about how such value measures are perceived, although

users of cost-benefit analyses appear not to have fully considered non-monetized

quantitative measures or qualitative assessments. In contrast, participative decision

making exercises have used ecological indicators (quantified but not monetized) as a

basis for valuation.

While there is little direct evidence on the perception of different kinds of
ecological value measures per sc\ other research on perception suggests conclusions
relevant for effective communication of ecological values. Response scales tend to
promote responses congruent with their structure. So, for example, asking people for
ecological value in dollars will likely elicit those values that are most readily expressed in
dollars, and not those that are difficult to express in dollars. However, numerical
information alone provokes weak - if any - affect, and is unlikely to influence
respondents' estimates of the value of the stimulus much (e.g. Dunn and Ashton-James,
2007, On emotional innumeracy), as is also demonstrated by studies on scope
insensitivity. Visual information often dominates other representations. Taken together,
this evidence suggests that quantitative cost benefit analyses will inevitably be more
strongly influenced by monetized values than qualitative or non-monetized quantitative
information that is not readily included in a cost benefit calculus. It also suggests that
attitudes, opinions, and values elicited based on qualitative and visual stimuli will
dominate those elicited based on numbers alone.

One mechanism for mitigating these disconnects related to ecological values
reported in different metrics is to employ an iterative, interactive approach to eliciting,
studying, and communicating values and tradeoffs, in which values are represented in
multiple ways. To exemplify the potential pitfalls: verbal quantifiers (e.g., "many" or
"very likely") are often proposed as a way of making technical information more
accessible but the wide variability with which these terms are interpreted (Budescu and
Wallsten, 1995) makes it critical to make the underlying numerical information readily
available. Appropriate use of graphical and visual approaches including geographic
information systems can aid interpretation of quantitative information. Visualization can

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facilitate new insights (MacEachren, 1995).

Interactive communication of ecological valuation information is likely to be
more effective in many circumstances than static displays. Interactive communication
allows users to manipulate the data or representations of the data, such as with sliders on
interactive simulations. Interactive visualization has the potential to allow users to tailor
displays to reflect their individual differences and questions. Even with exactly the same
presentation, because of differences in educational or cultural background, and different
intellectual abilities, people's understandings of presented content vary. Interactive
exploration tools give the audience a chance to investigate freely the part that they are
either interested in or about which they still have questions.

As argued by Strecher, Greenwood, Wang, & Dumont (1999), the advantage of
interactivity lies in: a) allowance for active, instead of passive, participation of the
audience; b) the ability to tailor information for individual users; c) the ability to assist
the assessment process; and d) the ability to visualize possible risks under different
hypothesized conditions (allow users to ask "what if' questions). Interactivity is a good
solution if the complexity of the visualization has the potential to overwhelm users
(Cliburn, Feddema, Miller, and Slocum 2002). Interactive visualization nonetheless
poses challenges as well. Interactivity is necessitated and challenged (by the sheer
computational power required) at the same time by 3-D visualization, which has become
increasingly popular in visualization practice (Encarnacao et al. 1994).

In order to assess how much confidence to attribute to the projections involved in
the valuation, decision makers must be informed about the analyst's own judgment of the
uncertainty of the valuation and its prior steps, and the assumptions underlying the
valuation analysis. Making decision makers aware of these assumptions is important
because decision makers often have to explain and justify their decisions by clarifying the
assumptions driving the analysis.

5.2.3. Communicating Uncertainties and Ecological Valuation

Finally, because ecological valuation involves multiple kinds of uncertainty,

effective communication regarding ecological valuation involves effective
communication of uncertainties both to decision makers and to the public. In order to
convey to decision makers the degree of uncertainty in an ecological valuation, the

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simplest expressions - whether quantitative (measures of dispersion, such as variance) or

qualitative (such terms as "likely," "very likely," etc.) - are typically inadequate.

Analysts can specify the central tendency of an estimate (mean or median value, as

appropriate) plus a confidence interval (for example, the 95% confidence interval around

the mean value, or the range of estimated values), but in some cases this may require

possibly arbitrary judgments on the part of the analyst (Moss & Schneider 2000).

Furthermore, providing decision makers with such ranges of results can be highly

misleading, because those without training in probability and statistics might be led to

faulty assumptions, such as that the probability distribution of values between the end-

points is uniform. Sensitivity analysis can help in this regard, although what is really

needed is a description - verbal or pictorial - of the full probability distribution.

Institutional obstacles to conveying uncertainty may be related to the
understandable reluctance of analysts to expose themselves and their work to the risk of
appearing to be lacking in rigor. Analysts may thus have an unfortunate incentive to
exclude or otherwise downplay components of their analyses that they fear may
jeopardize the credibility of their overall effort. Suppressing less certain information runs
counter to the need for transparency and the reality that all estimates have some degree of
uncertainty (Arrow et al. 1996).

Historically, efforts to address uncertainty in ecological valuations - and more
broadly, in economic benefit assessments that are part of Regulatory Impact Analyses -
have been limited. But guidance set forth in the U.S. OMB Circular A-4 on Regulatory
Analysis in 2003 has the potential to enhance the information provided in RIAs regarding
uncertainty.

In the past, point estimates have been given far greater prominence in public
documents such as RIAs and other government valuations than discussions of uncertainty
associated with them. Uncertainty assessments are often relegated to appendices and
discussed in a manner that makes it difficult for readers to discern their significance.

This result is perhaps inevitable given that single point estimates can be communicated
more easily than lengthy qualitative assessments of uncertainty or a series of sensitivity
analyses. The ability of Monte Carlo analysis to produce quantitative probability
distributions provides a means of summarizing uncertainty that can be communicated
nearly as concisely as point estimates. The need for and means of communicating

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uncertainty in such a fashion has been addressed in the existing literature. If a summary

of uncertainty in an estimate is not given prominence relative to the estimate itself,

context for interpreting the estimate and opportunities to learn from uncertainty

associated with it may be lost.

Some resistance to the use of formal uncertainty assessments such as through
Monte Carlo analysis, and prominent presentation of the results may be due to the
perception that such analysis requires more expert judgment and therefore renders the
results more speculative. Also, some might argue that, given the inevitably incomplete
nature of any uncertainty analysis, prominently presenting its results would incorrectly
lead readers to conclude that the results of an ecological valuation are more certain than
they actually are. Both concerns seem to be unfounded. First, as described above,
developing characterizations of uncertainty (such as for inputs in a Monte Carlo analysis)
often simply involves making explicit and transparent those expert judgments that
necessarily already must be made to develop point estimates for those inputs. Moreover,
to the extent that an uncertainty analysis is thought to be incomplete in its
characterization of uncertainty, that fact can surely be communicated qualitatively.
Finally, MacEachren et al. (2005) suggest animation as an effective technique for
conveying uncertainties in space-time processes, which can help viewers distinguish
between spatial and temporal uncertainties. It's important to communicate uncertainty
appropriately in all contexts, regardless of the difficulty of doing so.

5.2.4. Recommendations

In conclusion, the committee provides the following preliminary recommendations to

assist EPA in strengthening the communication of ecological valuation information.

•	use the iterative approach described in this report where technical experts and
interested and affected publics interact to clarify the values to be represented in
the analysis to provide a foundation for effective communications

•	develop an empirical analysis of the users of valuation and adapt valuation
communication to their needs

•	follow demonstrably effective basic practices for risk and technical
communication

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•	evaluate communications to assess their effects and to learn how to improve upon

them.

•	use GIS and interactive geospatial information systems integrated with other
ecological models where feasible, to represent the state of ecological systems and
services. Use best cartographic principles and practices

•	use interactive tools for communications, where feasible.

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6 APPLYING THE APPROACH IN THREE EPA DECISION

CONTEXTS

This chapter discusses a number of important issues that arise in implementing the
ecological valuation approach as they arise in three specific EPA decision contexts, national
rule making, site-specific decision making, and regional partnerships. The committee
believes that improved ecological valuation in each context can contribute to improved
policy analysis and decisions. The committee examined a number of illustrative examples
for each decision context and used these examples to inform its views about application of
the approach advocated in this report.

The discussions in the sections below elaborate on the three key features of the
valuation approach advocated in this report as they relate to the specific decision contexts:
early identification of and focus on impacts that are likely to be most important to people,
predicting ecological changes in value relevant terms, and the use of multiple methods in the
valuation process. The discussions are meant to be illustrative rather than comprehensive.
For example, the exclusion of a particular method from discussion in a specific context is not
intended to suggest inappropriateness. Note that the general principles and concepts used in
the discussions below are described in more detail elsewhere in this report (see, for example,
Chapter 4 and Appendix B for descriptions of individual methods).

6.1. VALUATION FOR NATIONAL RULEMAKING

6.1.1. Introduction

The objective of this section is to examine the valuation of ecosystem services by the
Agency with an emphasis on the monetary valuation of the economic benefits and costs of
national rules promulgated by the Agency and to make recommendations as to how the C-
VPESS valuation framework could be implemented in this context.

Most of the environmental laws administered by the Agency require that regulations
such as environmental quality standards and emissions standards be based on a set of criteria
other than economic benefits and costs. Indeed, in some cases the legislation explicitly
precludes consideration of costs or economic benefits in the standard setting process. For
example, in the case of the Clean Air Act, rules to establish primary ambient air quality
standards for criteria air pollutants are to be set to protect human health with an adequate
margin of safety. Even in those cases where the law allows consideration of the economic

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benefits and costs, such as the Safe Drinking Water Act, adherence to a strict "benefits must

exceed costs" criterion is not required.

Nonetheless, an assessment of the economic benefits and costs of EPA actions plays
an important role in the context of national rule making for a number of reasons. First,
analyses of major Agency rule makings are required under the terms of Executive Order
12866 (as amended by Executive Order 13422), which states, "Each Agency shall assess both
the costs and the benefits of the intended regulations, and ... propose or adopt a regulation
only upon a reasoned determination that the benefits of the intended regulation justify its
costs" (Executive Order 12866, October 4, 1993). These assessments are commonly referred
to as regulatory impact assessments or RIAs. They generally evaluate in economic terms the
form and stringency of the rules that are established to meet some other objective such as
protection of human health. Second, an assessment of economic benefits and costs can be
mandated by law. For example, the prospective analysis of the economic benefits of the
Clean Air Act Amendments of 1990 was mandated by Section 812 of the Amendments,
which requires the Agency to develop periodic reports to Congress that estimate the
economic benefits and costs of various provisions of the Clean Air Act. Finally, the
economic benefit and cost estimates developed in national rule making may later be taken
into account by executive branch officials and legislators in formulating and proposing new
national rules or for other purposes. Therefore, a complete, accurate, and credible analysis of
the economic benefits and costs of a given rule can have broad impacts even if the analysis
does not determine whether the current rule is enacted.

Circular A-4 from the Office of Management and Budget (OMB, 2003) makes it clear

that what is intended by Executive Order 12866 is an economic analysis of the benefits and

costs of the proposed rules conducted in accordance with the methods and procedures of

standard welfare economics. Thus, in the context of national rule making, the terms "benefit"

and "cost" have specific meanings. To the extent possible, the economic benefits associated

with changes in goods and services or prices due to the rule are to be measured by the sum of

the individuals' willingness to pay for them. Similarly, the costs associated with regulatory

action are to be evaluated as the losses experienced by people and measured as the sum of

their willingness to accept compensation for those losses. Thus, the analysis begins by

specifically describing environmental conditions in affected areas, both with and without the

rule. These changes are then evaluated based on individual willingness to pay and to accept

compensation and aggregated over the people (or households) experiencing them. Circular

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A-4 includes recognition that it might not be possible to express all benefits and costs in

monetary terms. In these cases, it calls for measurement of these effects in biophysical

terms. If that is not possible, there should still be a qualitative description of the benefits and

costs (OMB, 2003, p. 10). While Circular A-4 is clear about what should be included in

regulatory analyses, it does not preclude the inclusion of information drawn from non-

economic valuation methods. We believe that including this information along with

economic estimates of benefits and costs can prove useful to decision makers in many

circumstances.

This section considers ecological valuation in the context of national rule making
governed by Executive Order 12866, as amended, and OMB's Circular A-4. It focuses on the
use of economic valuation methods that seek to monetize economic benefits based on the
concept of willingness to pay (or accept compensation), recognizing that when monetization
is not possible, the Agency should seek to quantify impacts in biophysical terms or provide a
science-based, qualitative description as required by Circular A-4. As background for this
discussion, the committee examined three specific examples of previous Agency economic
benefits assessments: a) the Agency's benefit assessment for the final effluent guidelines for
the aquaculture or the concentrated aquatic animal production industry (US EPA 2004); b) its
assessment for the recent rule making regarding concentrated animal feeding operations
(CAFOs) (US EPA 2002; see also the discussion in Chapter 2 of this report); and c) the
prospective analysis of the benefits of the Clean Air Act Amendments of 1990 (US EPA
1999).36 Brief descriptions of the three benefit analyses are presented in separate text boxes.
These examples provide insights that are reflected in the discussion and recommendations
throughout this section.

6.1.2. Implementing the Proposed Approach

This section describes how EPA could implement the integrated and expanded

approach to ecological valuation proposed in this report in the context of national rule

making and RIAs. It illustrates how the three key features of the C-VPESS approach could

be implemented in this context. .

6.1.2.1 Early identification and focus on socially important impacts

Identification of the socially important impacts of a given rule requires information

about both the potential biophysical effects of the Agency's actions and the ecological

services that matter to people. To guide the collection of this information, the Agency should

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develop a conceptual model of the ecological and economic system being analyzed.

Conceptual models can allow the Agency to take a broad view of the complexities involved

in addressing ecological changes (see discussion in 6.1.2.2). It should be standard practice

for the Agency to develop such a conceptual model before other analytical work begins on an

economic benefit assessment or RIA. The analytical blueprint required as part of EPA's

process for developing rules should call for development of a conceptual model for

ecological valuation and specify the interdisciplinary team to be involved in developing it.

Determination of the relevant ecological effects to include in the conceptual model
could draw on technical studies of impacts and their magnitudes, as well as solicitation of
expert opinion regarding the nature of physical and biological effects of a regulatory change.
As an example, Figure 6 gives a general overview of the ecological impacts of CAFOs,
which enables a comprehensive evaluation of what is happening to the environment and
where the levers are for improving environmental performance. Although the CAFO rule
was adopted pursuant to the Clean Water Act, RIAs do not need to restrict the benefit
measure to the direct focus of the authorizing statute. As illustrated in Figure 6, the
environmental impacts of CAFOs extend beyond simply the water quality impacts. For
example, CAFOs are the source of interactive pollutants that impact the air as well as the
water. Further, the feed supply chain providing inputs to CAFOs involves many adverse
environmental impacts that need to be considered. A comprehensive overview such as this
could be used to develop a conceptual model that identifies potential ecological services that
might be affected by CAFO regulation.

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Figure 6: General Overview of the Impact of CAFOs

\

A.

Nitrate

Phosphate

Sediments

Hormones

Antibiotics

Pesticides

Heavy Metals

Pathogens

Salts

Organic Matter

The conceptual model should reflect not only the ecological science but should also
be based on information about the changes that are likely to be of greatest importance to
people. In the past, the Agency has generally chosen to focus on impacts that can be
monetized with readily available techniques or estimates from the existing literature, or both.
All three of the rule making benefit assessments that the committee reviewed provide
evidence of this practice. For example, for both the CAFO and the aquaculture rules, the
focus of the assessments appears to have been driven largely by the ability to use existing
estimates of willingness to pay for water quality improvements taken from Carson and
Mitchell contingent valuation study that had been used in previous EPA rule makings
(Carson & Mitchell 1993). Rather than choosing the focus based on ability to monetize, the
Agency should seek to identify those impacts that are likely to be of greatest importance to
society.

The committee believes that identification of socially important impacts reflecting

public preferences cannot be done deductively. Rather, it requires an examination of the

evidence regarding public preferences, i.e., what matters to people. This can be gleaned from

a variety of research approaches. In considering alternative approaches, it is important to

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distinguish the research approaches that can provide valuable information about the goods

and services that are important to people from the approaches to be used to evaluate

contributions to human well-being and costs. Where the analysis is being conducted to meet

a mandate for economic benefit-cost analysis (as is the case for RIAs), the computation of

economic benefits and costs must be consistent with the methodological requirements of the

benefit-cost framework. However, the process of identifying early on the public concerns

associated with a given rule can be undertaken with a variety of methods.

The suite of methods that can be used to identify the socially important ecological
changes includes surveys, public meetings, focus groups, and content analysis of public
comments. More specifically, possible approaches for obtaining information about public
preferences and concerns include:

•	Inventory of the reasons invoked in similar rule making processes in other
jurisdictions (e.g., state and local).

•	Inventory of the concerns expressed in public hearings at various
governmental levels (perhaps with weightings based on the frequency of
concerns raised). For example, local vs. national concerns can be quantified
through content analysis of transcripts.

•	Focus groups and surveys of concerns (can be lists of concerns, or quantified
by ranking priorities).

Relevant initiatives, referenda, or community decisions might also be available in some
jurisdictions to get a more robust indication of the preferences for various types of ecosystem
services or the avoidance of the various risks.

An important consideration in identifying socially important impacts is the extent to
which the public understands the role that ecosystems play in providing services that
contribute to human well-being. Many ecosystem services, while well known to the
scientific community, are little known or misunderstood by the general public (Weslawski, et
al. 2004). For example, the full chain of connections in the production of animals in
CAFOs, as described in Figure 6, is not generally understood or appreciated by the public.
Similarly, the public does not generally understand the organisms and processes involved in
breaking down waste products, or the services provided by those processes. For example,

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certain groups of soil organisms maintain soil structure by their burrowing activities, while

others facilitate the use of nutrients by higher plants for growth. This problem of lack of

public understanding might be exacerbated in national level analyses where ecological

impacts and vulnerabilities can vary substantially across locations. When relying on

information from public expressions of preferences (e.g., surveys, public hearings,

community decisions) to identify socially important impacts, the Agency should assess

whether, when expressing preferences, the public understood these contributions sufficiently

well to provide informed responses.

The above discussion envisions a process under which the analysts conducting the
valuation would develop the conceptual model, drawing on information provided by other
experts and the public. Alternatively, the conceptual model could be developed through a
more participatory process, such as mediated modeling (see Appendix B). Participatory
processes can be particularly useful when the services generated by an ecosystem are not
well-understood by the public and hence information about public preferences expressed
through non-participatory methods may be misleading. While time and resource constraints
may preclude use of such a process in many contexts, the committee suggests that EPA
experiment with participatory processes (for example, holding open meetings for the public
and Agency staff) to aid in the development of the conceptual model for a particular rule
making. Such an approach would provide an interactive forum for determining the
ecological changes that are important both biophysically and socially.

Regardless of the specific process used to develop the conceptual model and identify
the ecological impacts that will be the focus of the valuation exercise, in order to increase
transparency the Agency should document in its economic benefit assessments and RIAs
how the decisions underlying the conceptual model were made. It should clearly identify the
criteria for including effects within the core analysis and how these criteria were applied to
those analytical choices. In addition, EPA should specifically document in final economic
benefit assessments and RIAs how the Agency incorporated relevant input on ecological
values related to the rule from public meetings on the proposed rule. It would also be helpful
to provide a specific section in RIAs and economic benefit assessments describing how the
Agency addressed the most significant public comments regarding ecological values and
valuation. Finally, the final conceptual model that was used to guide the analysis should be
part of the public record for every rule making and be available online.

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6.1.2.2 Predicting biophysical changes in value-relevant terms

As discussed in Chapter 3, the C-VPESS approach calls for the use of ecological

production functions to operationalize the conceptual model. This requires first a prediction

of the change in relevant stressors resulting from an EPA action, and then a prediction of

how that change will affect the ecosystem and ultimately the provision of ecosystem services.

In some cases, the links between stressors and ecosystem services are well understood

and relatively easily quantified. Examples include the movement of phosphorus and nitrogen

from manure into surrounding waters. Phosphorus in particular has been studied intensively

and, importantly, its impact has been well demonstrated by whole ecosystem experiments for

fresh water.37 Similarly, species that the public or experts particularly value have been

studied in sufficient detail that there are process models of production and interaction with

other species. Scientists can specify an ecological production function for these organisms

and use that function to predict the impact of changes in stressors.

However, for many services, developing the relevant ecological production functions

is much more difficult, particularly in the context of national rule making, for a number of

reasons. First, in many rule making contexts, predicting the changes in stressors is difficult.

As illustrated in Figure 6, CAFO operations involve many stressors with complex

interactions, which greatly complicate the development of quantitative estimates of changes

in stressors. In addition, changes must be defined relative to a baseline, which might not be

known. For example, in the RIA for the aquaculture rule, it was difficult to quantify the

changes in stressors because in some cases baseline data on stressor levels were not

available. In other cases the rule only required best management practices rather than

quantitative maximum discharge levels, and it was difficult to predict how the adoption of

best management practices would affect the stressors.

Second, many of the links between stressors and ecosystem services are not fully

understood by scientists. For example, one of the important ecosystem services affected by

the CAFO rule is the support of populations of fish species that are targets of recreational

angling. To predict the effects of the rule on ecosystem services, one would need to know

how populations of these species change and how population changes affect anglers' success

rates. These links are not well understood at the level required for a comprehensive national

analysis. Scientific knowledge is especially lacking in understanding the ecological impacts

of substances such as heavy metals, hormones, antibiotics, and pesticides. Yet these

substances can have important and far-ranging impacts that could be significant at the

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national level. For example, arsenic in poultry manure moves into local environments as

well as through different pathways to places more distant, either through the sale of

incinerator ash for fertilizer or by the use of dried and pelletized manure (Nachman, et al.,

2005).

Finally, both the nature and magnitude of impacts can have substantial variation
across regions of the country, implying the need for a more comprehensive analysis. Yet
comprehensive analysis is particularly difficult precisely because of this scale and the
associated complexity. For example, the committee's review of the CAFO rulemaking noted
the following issues that stem from the varied and complex environmental consequences of
CAFOs (see Figure 6):

•	Multimedia effects, i.e., interrelated impacts on both water and air quality;38

•	Impacts across multiple geographical scales (e.g., local, regional, global);39

•	Differences in the time persistence of pollutants (e.g., days vs. decades);40

•	Geographical clustering and the need for site-specific analysis due to
uniqueness of site characteristics associated with impacts;41 and

•	Ecological impacts through supply-chain effects that are geographically
dispersed.42

Thus, the combination of variation, complexity, and gaps in information and
understanding make it difficult for the Agency to assess the ecological impacts of its actions,
particularly at the national scale. Yet, this is an essential component of benefit assessment
for national rule making, as laid out in Circular A-4. As noted previously, Circular A-4
requires the Agency to monetize impacts that can be monetized, quantify those that cannot be
monetized but can be quantified, and describe qualitatively (based on scientifically-credible
theories or evidence) impacts that cannot be quantified. Despite the difficulties described
above, the approaches to predicting ecological impacts discussed in Chapter 3 can help the
Agency meet the requirements of the Circular, through providing scientifically-based
qualitative descriptions of ecological impacts and then, where possible, quantifying those
impacts for use either as a means of quantitatively describing effects that cannot be
monetized or as an input into monetization of the associated values.

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As noted above, characterization of ecological impacts requires a conceptual model

(see detailed discussion in Chapter 3, Section 3). Such a model would link the various levels
of organizations of ecosystems that are involved in the provision of ecosystem services, as
illustrated in Figure 6. A carefully developed and scientifically-based conceptual model can
be used as the basis for a qualitative but detailed description of the ecological impacts of a
given change. A listing that simply summarizes possible impacts, however, is insufficient.
Such a summary should be accompanied by justification based on the conceptual model and
the associated theoretical and empirical scientific literature. To the extent possible, the
existing literature should be used to draw inferences about the likely magnitude or
importance of different effects, even if only qualitatively (e.g., high, medium, low).

To move from a qualitative to a quantitative prediction of impacts, the impact of
changes in stressors on the ecological system must be estimated, and the predicted changes in
the ecological system must then be used to quantify predicted changes in the provision of
ecosystem services. To do this, the conceptual model must be linked with one or more
ecological models that capture the essential linkages embodied in the conceptual model and
are parameterized to reflect the range of relevant scales and regions. The objective is to use
the models to generate metrics to compare biological conditions with and without the rule to
see the potential effect of the rule on the delivery of ecosystem services. Since there may be
a long chain of ecological interactions between the stressors and the ecosystem services of
interest, the use of quantitative models of the various components of the system will often be
required to determine the net effect of these interactions on the levels of ecosystem services
of concern. Outputs from these models give quantitative values of the stressor impacts even
though all cannot be monetized.

Ecological models are currently utilized in rule making. However, sometimes their

complexity, cost, and time constraints encourage the use of the simplest modeling approaches

available that can be tailored to economic valuation. In addition, as noted previously, in the

past the Agency has generally chosen to focus on impacts that can be monetized with readily

available techniques and chosen ecological models based on this rather than on the important

links identified in the conceptual model. For example, for the aquaculture rule, the Agency

used the QUAL2E model to predict ecological impacts. While this model can estimate the

interactions among nutrients, algal growth, and dissolved oxygen, it is not capable of

ascertaining the impacts of total suspended solids, metals, or organics on the benthos and the

resulting cascading effects on aquatic communities. The choice of QUAL2E appears to have

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been driven largely by the ability to link its outputs with existing estimates of willingness to

pay for water quality improvements taken from Carson and Mitchell contingent valuation

study that had been used in previous EPA rule makings (Carson & Mitchell 1993).

Chapter 3 Section 2 discussed the need to develop criteria for choosing among
alternative models. In general, rather than basing this choice on the ability to link to existing
value estimates, the Agency should use the conceptual model of the relevant ecosystem(s) to
guide the selection of ecological models and seek to predict the impacts of changes in
stressors on a broader set of potentially important ecosystem services.

Chapter 3 also discussed the use of indicators when available ecological models do
not provide a full characterization of the relationship between ecosystem structure and
function and the provision of ecosystem services. For example, fully tested techniques are
available for evaluating different functional groups, and, in theory, metrics related to these
groups could be used to quantify the ecological impacts of a given rule (see Figure 5).
Specifically, the abundance of these groupings can be readily quantified in any before-and-
after rule condition. For example, at the base of the ecosystem is its potential and realized
biological diversity. Thus, metrics that look at the impact of a rule on species richness and
various diversity indices can quantify potential and realized biological diversity Such
metrics, however, cannot be tied directly to the provided ecosystem services without
embedding this information into an ecosystem model that reveals ecological functions and
related services. The key, though, is to identify those components of each of the functional
levels that are most directly related to the services of interest and thus provide ecological
indicators of the state of the system in relation to the change in stress level. A number of
approaches are able to limit the indicators to those that will provide the most direct
information relevant to the services in question. One approach is to focus on those functional
groups that play a most prominent role in service provision as noted above.

Finally, the site-specific nature of many ecological impacts makes national level

benefit assessments difficult. This difficulty has been noted and discussed by the SAB in

previous advisories, including the Advisory on EPA's SuperfundBenefits Analysis (2006d).

Rather than conducting a "top-down" analysis at the national level, to address variability

across sites the Agency should explore the use of a "bottom-up" approach. Under this

approach, a number of case studies that reflect different types of ecosystems could be

conducted. If information about the distribution of impacted ecosystem types is available,

these case studies could in principle be aggregated to provide national level estimates of

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impacts. Even without full information about the distribution of ecosystem types affected by

the rule, the individual case studies could still provide information about the range of impacts

and their dependence on ecosystem characteristics. This information could be useful not

only for the specific policy decision for which it was conducted, but also in guiding future

research. In particular, it could suggest key ecosystem characteristics that would be useful in

categorizing ecosystems for future valuation analyses, and for which additional information

about their distribution is needed.

In summary, the initial conceptual model of a system, which provides the big picture
of the possible environmental impacts of the rule, can provide a detailed and scientifically-
based way of qualitatively characterizing the ecological impacts of a rule. Even if some of
the identified effects cannot be quantified, this detailed characterization will provide valuable
information regarding the impact of the rule. Ecological models can then be used to
operationalize the conceptual model and quantify impacts, where possible. The choice of
models should be guided by the conceptual model rather than by the ability to easily
monetize the model's outputs. The quantification should consider not only changes in a set
of final ecosystem services (e.g., clean water), but also changes in intermediate services
when the contribution of those services is not fully captured by the final services included in
the assessment. Even when changes in ecosystem services cannot be quantified explicitly,
metrics can be used that would indicate the success of rule making in providing better
ecosystem services to society. These can provide a means of quantifying impacts that cannot
be monetized, and, where feasible, serve as an input into monetizing or otherwise
characterizing the value of the changes in ecosystem services. In addition, site-specific
variability can be addressed by including in the benefit assessment case studies for important
ecosystem types, with the possibility of aggregating across these case studies if information
about the distribution of ecosystem types is available.

6.1.2.3 Monetary Measures of Value

To comply with the requirements of Executive Order 12866, as amended, Circular A-

4 calls for the monetization of economic benefits whenever possible. Although a variety of
methods can be used to determine values for purposes of identifying socially relevant
ecosystem characteristics and services (see discussion in Chapter 6 Section 1.2.1) and for
value assessments in other contexts (see Chapter 6, Sections 2 and 3), in the context of
economic benefit-cost analysis the only approach to monetization consistent with the

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premises underlying this analysis is the use of economic valuation methods. Monetizing

values using other methods and then aggregating the resulting estimates is problematic

because it implies adding together numbers that are based on quite different methods and

underlying premises. Thus, for both theoretical and empirical consistency, the monetization

of benefits in a benefit-cost analysis should be based on economic valuation.

Economic valuation methods are well-developed and there is a large literature
demonstrating their application. (See Chapter 4 and Appendix B for descriptions of
economic valuation methods.) Nonetheless, applying these methods to national-level
analyses of the ecological benefits of a rule is difficult. As with the prediction of ecological
impacts, the value of ecological impacts is likely to be site-specific, depending on local
conditions and the characteristics of the affected population. As a result, generalizing to the
national level is difficult. In principle, this variability across affected sites could be
addressed by conducting case studies and aggregating the results across the sites affected by
the rule. However, time and resource constraints may preclude doing this kind of original
economic benefits research. As a result, the Agency will generally need to rely on benefits
transfer instead. Although the existing economic valuation literature is extensive, most of the
previous ecological valuation studies that might serve as study sites are not national in scope.
Rather, they involve valuing relatively localized changes affecting a local or regional
population. In addition, these studies have generally focused on a limited number of
ecosystem characteristics or services (primarily related to recreation). Few studies provide
national level value estimates for a range of services that could be readily used in a national
level benefit assessment, [is this an accurate statement? I think so, but need to check - KS]

The Agency needs to ensure that the call for monetization, coupled with the need to

use benefits transfer and generate national-level benefit estimates, does not unduly restrict

the types of ecosystem impacts considered in the economic benefit assessment, or lead to

inappropriate application of economic valuation methods or benefits transfer. As noted

above, in the past Agency decisions regarding the focus of ecological benefit assessments

have been driven to a large extent by the objective of monetizing the value of impacts at the

national level using benefits transfer. This applies both to the types of ecosystem services

included in the detailed assessments and to the choice of the ecological models used to

predict biophysical impacts. For example, the Agency's assessment of the CAFO rule

focused primarily on valuing recreational impacts, driven to a large extent by the ability to

link the QUAL2E model with off-the-shelf monetary estimates of willingness to pay for

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changes in water quality indices taken from the Carson-Mitchell contingent valuation (CV)

study. The principal advantage of this approach is that it utilizes a study designed to be

national in scope and has a simple willingness-to-pay relationship that allows the analysis to

be done relatively quickly, without new research and the associated significant expenditures

on research resources. In addition, it can be applied using a straightforward conceptual logic

that is easy to understand. However, use of the Carson-Mitchell estimates has a number of

limitations that raise concerns about the resulting economic benefit estimates. Most notably,

the study was conducted more than 20 years ago. In addition, it was designed for a different

purpose and was not intended to apply to specific rivers or lakes. The water quality index

was highly simplified and was never designed to reflect ecological services related to water

quality (other than those related to fish). Thus, in an effort to focus on effects that could be

readily monetized at the national level using benefits transfer, the Agency appears to have

limited both the types of services considered and the ecological and economic models used to

estimate the impacts of the rule on those services.

Since the Agency will inevitably need to rely on benefits transfer for many, if not
most, RIAs, it must take care to ensure that the transfer of economic benefit estimates is
appropriate. Chapter 4 discusses issues that arise in transferring economic benefit estimates
and provides suggestions for ensuring that the transfer is appropriate, given both the
biophysical and the socio-economic characteristics of the study and policy sites. The use of
the Carson-Mitchell study to estimate the benefits associated with the proposed CAFO rule
provides an example where the transfer of benefits was problematic. However, in other cases
EPA has appropriately used benefit transfer. For example, EPA estimated the recreational
benefits of reducing acid deposition in Adirondack's lakes by transferring benefit estimates
from a fairly recent published study of recreational angling choices of households in New
York, New Hampshire, Maine, and Vermont (Montgomery and Needelman 1997; for more
detail see Text Box 4: The Prospective Economic Benefits of the Clean Air Act
Amendments). This study explicitly compared populations of target species and pH levels at
the source and target sites. If the socio-economic characteristics of the population of these
four New England states match those of the Adirondacks region of New York State, this
study is a good source for economic benefits transfer.

The above example illustrates a benefits transfer based on an individual RUM study.

As discussed in Chapter 4, benefits transfer can also be based on meta-analyses, which

combine information about values from multiple studies. For example, several studies have

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used random utility models to link physical descriptors of water quality to recreation

behavior to estimate the willingness to pay or willingness to accept per recreational trip for a

change in water quality (e.g., water quality that improves to boatable or fishable status), had

it been experienced in each of the areas. These estimates could be used in a summary or

meta function describing how the local choice set of recreation sites and economic

characteristics of the recreationists, as well as the character of the changes from existing

baseline conditions, influenced the estimates of unit economic benefits. If the changes

considered in these studies are comparable to what would have been experienced under the

proposed rule, then the meta function could be used to estimate values at sites affected by the

rule, (references)

Alternatively, the models could be adapted to be directly applied to choice sets
composed for affected areas. In this case the recreation behavior necessary to operationalize
the model could be extracted for some of the areas from EPA's National Survey on
Recreation and the Environment (NSRE) for 2000 and 2004. The logic involved has two key
steps: a) translation of the effect of the rule for a set of local water quality conditions that is
matched to some set of economic behavior for that area that is influenced by the water
quality; and b) adaptation of an economic model of trade-offs people would be willing to
make to improve one or more aspects of the water quality for the area so that economic and
ecological factors affecting the trade-offs are represented in the summary function. There is
precedent in the literature on economic benefits transfer for these types of analyses (see
Rosenberger and Loomis 2003 and Navrud (in press), for examples of how this logic might
be used in benefits transfer). [I don't understand the idea behind this second approach from
the description here. What is the key distinction? I think it would be helpful to have some
clarification, but I can't revise this to be clearer without more info. KS]

A second class of studies for transferring benefits using meta analyses are the stated

preference and stated choice studies (such as Carson and Mitchell) that highlight water

quality attributes. While the record here is not as extensive as it is for the revealed

preference random utility studies (RUMs), there are several candidate studies (references??).

These analyses are based on surveys that require respondents to choose from among a set of

options, such as plans for reducing effluents or improving water quality. The logic is

comparable to that described for the RUM. The effects of the rule need to be adapted to the

features of each of the models, and the projected unit economic benefits must be derived.

Then the factors affecting the economic benefit measure for each are modeled in a summary

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analysis that can be applied to other areas that are affected by the rule. [It is not clear to me

whether this paragraph was intended to refer to meta-analyses based on stated preference, or

to the transfer of benefit estimates from individual SP studies (like Carson and Mitchell). It

looks like the former, but I'm not sure. We should try to clarify this. KS]

As noted above, the existing literature on economic valuation of ecosystem services

has focused to a large extent on estimating the value of recreational impacts. In addition to

recreational impacts, some ecological services affect the well-being of homeowners living

near the ecological systems providing these services. Examples include water regulation and

flood control, and the amenities associated with healthy populations of plants and animals.

The willingness of residents to pay for these services can be capitalized into housing prices.

The hedonic property value method can be used to obtain estimates of the values of these

services. Examples illustrating this approach to valuing ecosystem services include Leggett

and Bockstael (2000), Mahan, et al. (2000), Netusil (2005), and Poulos et al. (2002).

Estimates from studies such as these could be candidates for use in an economic benefit

transfer. However, as with the recreation studies, these studies tend to be local rather than

national in scope, which makes extrapolation to national level benefit assessment difficult.

The above discussion suggests that, to improve the Agency's ability to value the

ecological impacts of national rules using economic valuation methods, additional research is

needed to a) generate national-level value estimates that can be used in benefits transfer,

particularly for recurring rulemakings, b) generate information about the distribution of

ecosystem and population characteristics across local or regional sites that could be used to

aggregate localized case studies in a "bottom-up" approach to national-level analysis, and c)

expand the range of ecosystem services valued using economic valuation methods so that

benefits transfers can incorporate a wider range of services.

6.1.2.4 The Role of Other Valuation Methods

Although Circular A-4 calls for the use of economic valuation methods to monetize

benefits, other valuation methods can also play an important role in RIAs. The valuation

approach proposed in this report envisions three possible roles for other valuation methods in

the context of national rule making.

First, as already discussed, other methods can be used to identify early in the process

those ecosystem characteristics or services that are likely to be socially important and hence

should be a focus of the analysis. For example, focus groups, participatory/interactive

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processes, surveys of attitudes and judgments, analyses of public views regarding related

ecological impacts (expressed through hearings, public comments, citizen juries, etc.) and

other similar methods can provide valuable public input into the development of a conceptual

model that captures the most important ecosystem services.

Second, although in principle economic valuation methods can fully capture the
relevant population's willingness to pay for changes in ecological systems and services, as
discussed above, in practice there are significant limitations that can make this very difficult,
particularly at the national level. When benefits cannot be monetized, Circular A-4 calls for
them to be quantified, or at least qualitatively described, to the extent possible, using
scientifically-based analysis. When estimates of willingness to pay cannot be generated for
the full range of important ecosystem services, it may be possible to use other methods as
proxies for, or indicators of, willingness to pay. To the extent that other methods generate
non-monetary measures that are likely to be correlated with willingness to pay, they can
provide useful information about likely changes in willingness to pay when direct monetary
measures of those changes are not available. For example, economic benefit indicators (see
Chapter 4 and Appendix B) can be viewed as non-monetary measures of impacts that are
likely to be correlated with willingness to pay; ceteris paribus, the more people living within
the vicinity of an impacted ecosystem, the higher is the willingness to pay to protect that
system likely to be. Similarly, ceteris paribus, the more people who judge the protection of a
given ecosystem service to be "somewhat important" or "very important" in a survey of
attitudes and judgments, the higher is the willingness to protect that service likely to be.
While use of these proxies would not provide monetary estimates of benefits that could be
compared to cost, they can provide important information about possible benefits. Care must
be taken, however, to avoid misinterpretation of these proxies. For example, just because a
large population lives in the vicinity of an impacted ecosystem does not necessarily mean
that a change in that ecosystem has a large value. If the change relates to a service that is not
important to people, the value of that change (i.e., the willingness to pay for it) would be low
regardless of the number of people living in the vicinity. To draw correct inferences, the
Agency would need information not only about the number of people affected but also about
the importance that individuals attach to the service, as revealed for example through surveys
or other methods.

Finally, although benefit-cost analysis requires the use of economic valuation to

estimate benefits, RIAs need not be limited to information generated for use in a strict

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comparison of benefits and costs. Information about other sources of value that do not fit

within the theoretical framework underlying benefit-cost analysis can still be useful to policy

makers when making decisions related to ecological protection. For example, the religious,

spiritual, or cultural value of some ecosystems and their related services may be an important

consideration not adequately captured by standard measures of willingness to pay. Valuation

methods other than economic valuation can provide information about these other sources of

value. However, as noted previously, even when other methods yield value estimates

measured in monetary units (dollars), these values should not be added to monetary estimates

derived from economic valuation methods, since they are not based on the same underlying

assumptions and principles.

6.1.2.5 Reporting Value Estimates

To assess and report on changes in service flows, economic benefit assessments and

RIAs should feature prominent discussions of ecological services that describe how these
services were identified and analytical choices were made. In addition, consistent with the
guidance in Circular A-4, they should clearly identify the values that were a) monetized
using economic valuation methods; b) quantified (but not monetized); and c) described
qualitatively. If methods other than economic valuation are used to provide non-monetary
quantitative or qualitative information about benefits, the RIA should include a discussion of
the extent to which they provide proxies for, or indicators of, willingness to pay (or accept).
If methods other than economic valuation are used to capture sources of value other than
those typically reflected in willingness to pay, the methods used and the results should be
described in a separate section of the RIA as supplemental information.

Rather than simply designating some impacts as "non-monetized," as in the CAFO
benefit assessment, the committee recommends that the quantified but non-monetized
impacts be reported explicitly (in conjunction with the monetized economic benefits). For
those described only in biophysical terms, they should also be measured in the units that
make sense from a biological perspective, and the non-quantifiable impacts should be
described in as much detail as is feasible. Furthermore, any summary listing of the economic
benefits and costs should include all three types of contributions to human welfare with the
monetized and quantified values measured in the appropriate units (dollars or biophysical
units). When monetized economic benefits are aggregated, the resulting sum should always
be described as the "Total Economic Monetized Benefits" rather than the "Total Benefits."

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In the past, EPA has sometimes reflected the non-monetized economic benefits in aggregate

measures of benefits by including an entry in the summary table of economic benefits (and

costs) such as +X or +B to indicate the unknown monetary value that should be added to

economic benefits if the value could be determined. While such an approach indicates that a

measure of monetary economic benefits (and costs, too, if appropriate) is incomplete, the +X

or +B designation provides insufficient information and can be easily overlooked when the

results of the economic benefit assessment are used. Always designating the sum as "Total

Monetized Economic Benefits" provides a continual reminder of what is (or is not) included

in this measure. By also including key quantified but non-monetized impacts that are

measured in biophysical units, the Agency will be providing a more accurate and complete

indication of total benefits, as called for by Circular A-4.

Because of the difficulties in estimating biophysical impacts of an EPA rule and the
associated economic benefits or costs that result from that rule, the Agency must characterize
the uncertainty associated with its assessment. EPA should include a separate chapter on
"Uncertainty Characterization" in each economic benefit assessment and RIA. This chapter
should discuss the scope of the economic benefit assessment, the different sources of
uncertainty [e.g., biophysical changes and their impacts, social information relevant to
values, valuation methods (including transfer of willingness-to-pay or willingness-to-accept
information)], and the methods used to evaluate uncertainty. At a minimum, the chapter
should report ranges of values and statistical information about the nature of uncertainty for
which data exist. For each type of uncertainty, information similar to that reported in the
Agency's prospective analysis of the economic benefits and costs of the Clean Air Act
Amendments (US EPA, 1999) should be reported and a summary of this information should
appear in the executive summary of the RIA or economic benefit assessment. Specifically,
EPA should report: a) potential source of error; b) the direction of potential bias for overall
monetary economic benefits estimate; and c) the likely significance relative to key
uncertainties in the overall monetary estimate.

6.1.3. Conclusions

To develop more comprehensive estimates of the ecological benefits associated with
national rules and regulations, the Agency needs a broader approach to ecological valuation
than it has typically used in the past. The expanded approach to valuation proposed in this

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report could be applied in the national rulemaking context. This would entail challenges, but

important opportunities for improvement as well.

To ensure that the benefit assessment considers all socially important impacts, the

Agency should develop a conceptual model of the ecological and economic system being

analyzed to serve as a guide or road map for the benefit assessment. Development of the

model requires input from ecologists, social scientists and the public. The conceptual model

should adopt a multimedia (air and water) perspective, since focusing on a single media

(such as water quality) can miss major interactions among media that impact ecosystem

services. In developing the conceptual model, the Agency should draw from research based

on a variety of different methods to determine early on in the process which of the possible

ecological impacts are likely to be of greatest importance to people and hence should be the

focus of the assessment. The committee recommends that the Agency consider use of an

open, interactive public forum for identifying issues of concern. In addition, it should

document in the RIA the process used to identify those ecosystem characteristics and

services that were included in the assessment, as well as those that were excluded.

Given a conceptual model, a significant challenge to ecological benefits assessment

for national rules is predicting how the levels of ecosystem services would be affected by the

rule, particularly at the national level. The combination of variation, complexity, and gaps in

information and understanding make it difficult for the Agency to assess the ecological

impacts of its actions, particularly at the national scale. Reasons for this include:

•	In some cases (e.g., requirements for best management practices, absence of baseline
data), the changes in the levels of ecological stressors are not known.

•	The models used in the analysis often do not predict changes in the relevant
ecosystem services. For example, the links between outputs of some ecological
models and human uses of the ecosystem are not known (e.g., the relationship
between changes in fish populations and changes in recreational angling).

•	The needed ecological data are often not available.

The Agency should take steps to improve its capacity for predicting the ecological
consequences of Agency policies and regulations at the national level. Possible steps include
developing better quantitative ecosystem models for predicting the consequences of changes

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in ecological stressors on the production of ecosystem services and developing better

baseline data on ecological stressors and ecosystem service flows. In addition, site-specific

variability can be addressed by including in the benefit assessment case studies for important

ecosystem types, with the possibility of aggregating across these case studies if information

about the distribution of ecosystem types is available. This bottom-up approach would

proceed by establishing separate estimates for each regional grouping of similar facilities and

then adding them together to obtain the national estimate.

Circular A-4, which serves as the Agency's guide for preparing RIAs, requires that
benefits be monetized, if possible, using economic values. Methods exist for estimating
economic values for at least some ecosystem services; these methods have been used to
estimate values in a number of cases. However, applying these methods to new cases
(including an expanded range of ecosystem services) to analyze proposed regulations at the
national level could require original research that is costly and time consuming. As a
consequence, the Agency will often have to resort to economic benefits transfers to estimate
ecosystem values for rule making. Since economic values are context dependent, steps must
be taken to ensure that the transfer of economic benefits information is appropriate. This will
very likely require a much larger set of value estimates than is currently available. The
Agency should continue to support research to build an improved database for economic
benefits transfer for ecosystem service valuation.

In the past, the Agency has selected the ecosystem services to include in its
assessment as well as the ecological models to use in quantifying impacts based on the
objective of monetizing benefits at the national level using off-the-shelf value estimates
(benefits transfer). This can lead to benefit estimates that are not scientifically sound.

Instead, the Agency should use the conceptual model to drive the choices about which
services to include, even if that choice implies an inability to monetize the associated values
at the national level. In cases where benefits cannot be monetized, Circular A-4 requires that
the impacts be quantified, if feasible, or qualitatively characterized. The conceptual model
can provide a detailed and scientifically-based way of qualitatively characterizing the
ecological impacts of a rule. Ecological models can then be used to quantify impacts, where
possible. The choice of models should be guided by the conceptual model rather than by the
ability to easily monetize the model's outputs. It might also be possible to use other non-
monetary valuation methods to develop metrics that would likely be strongly correlated with

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willingness to pay and hence serve as a proxy or indicator measure when monetized values

are not available.

To ensure that benefit assessments do not inappropriately focus only on those impacts
that have been monetized, EPA should report non-monetized ecological effects in appropriate
units in conjunction with monetized economic benefits. In addition, aggregate monetized
economic benefits should be labeled as "Total Monetized Economic Benefits" rather than
"Total Benefits." In addition, EPA should include a separate chapter on "Uncertainty
Characterization" in each economic benefit assessment and RIA.

Methods also exist for estimating non-economic values for at least some ecosystem
services. While these methods do not properly fit within a formal economic benefit-cost
analysis, they can provide important additional information to support decision making.

When value estimates from these methods are included in RIAs, the RIA should clearly both
the method and the results in a separate section.

In general, EPA should seek to build additional capacity, externally and in-house,
specifically designed to facilitate ecological valuation for national rulemaking, particularly
for recurring rule makings. The committee advises the Agency to develop an extramural
grant program focused on method development specifically for recurring rule makings (e.g.,
for rule making associated with programs like EPA's National Ambient Air Quality
Standards or Effluent Guideline programs). Such a focused effort could help develop
methods for expanded applications of monetary and non-monetary methods for valuing
ecological effects, which will assist Agency regulatory programs addressing ecological
protection issues.

The committee also advises the Agency to host annual Agency-wide meetings to
discuss methods used in regulatory impact analyses and economic benefits assessments, and
methods needed for full characterization of the effects addressed by the regulatory actions
associated with those efforts. One objective of this effort should be to build an improved
database for economic benefits transfer for ecosystem service valuation.

Text Box 2: The Aquaculture Effluent Guidelines

Title III of the Clean Water Act (CWA) gives EPA authority to issue effluent
guidelines that govern the setting of national standards for wastewater discharges to
surface waters and publicly owned treatment works (municipal sewage treatment

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plants) The standards are technology-hasccl. i e they are ha sod on the performance of
a\ ailahlc treatment and control technologies The proposed effluent guidelines for
the Concentrated Aquatic Animal Production Industry would require that all
applicable facilities pre\enl discharge of drugs and pesticides that ha\e been spilled
and minimize discharges of excess feed and de\elop a set of systems and procedures
to minimize or eliminate discharges of \ arious potential en\ ironmental stressors The
rule also includes additional qualitatix e requirements for flow through and
recirculating discharge facilities and lor open water system facilities (I S l-IW.

21)04)

I'or most of these requirements, it is not possible to specify the change in the I ex el s of
en\ ironmental stressors since the rule called lor adoption of "best management
practices" rather than imposing specific quanlilati\e maximum discharge I ex el s In
addition, for most of these stressors, baseline data on discharges in the absence of the
rule were not a\ ailable

The Agency identified the following potential ecological stressors solids: nutrients:
biochemical oxygen demand from feces and uneaten food, metals (from feed
additi\es. sanitation products, and machinery and equipment), food additi\es for
coloration, feed contaminants (mostly organochlorides). drugs, pesticides, pathogens,
and introduction of non-nati\e species Some of these (for example, drugs and
pathogens) were thought by the Agency to be \ery small in magnitude and not
requiring further analysis To this list. C-VPIiSS added habitat alteration from
changes in water flows

The Agency analyzed the effects of changes in these stressors on dissolxed oxygen,
biochemical oxygen demand, total suspended solids, and nutrients (nitrogen and
phosphorus) There appear to ha\e been two reasons why the remaining endpoints
were not quantified

•	The Agency lacked data on baseline stressor I ex el s and how regulation would
change these I ex el s

•	The Agency did not use a model capable of characterizing a wide range of
ecological effects The Agency used the ()l AI.2I- rather than the ax ailable
AQl ATOX model The choice of Ql A1.211 appears to liaxe been drixen largely
by the ability to link its outputs with the Carson and Mitchell x aluation model
described below

The Agency estimated benellts for recreational use of the waters and non-use x allies
To estimate these x allies, the Agency estimated changes in six water quality
parameters for 3<> mile stretches downstream from a set of representatixe facilities
and calculated changes in a water quality index for each facility The Agency then
used an estimated willingness-to-pay function for changes in this index taken from
Carson and Mitchell (ll^3) Carson and Mitchell had asked a national sample of
respondents to state their willingness to pay for changes in a water quality index that
would moxe the majority of water bodies in the I nited Slates from one lex el on a
water quality ladder to another, resulting in improx ements that would allow for
boating, fishing and swimming in successixe steps This contingent xaluation surxey
was conducted in llM2-X3 and was not intended to apply to specific rixers or lakes

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The aggregate willingness to pay for the change in the water quality index for each
representative facility was then used to extrapolate to the population of facilities of
each type affected by the rule.

Text Box 3: The CAFO Effluent Guidelines

Context:

In recent years there has been substantial growth of the livestock industry in the
United States as well as in many other parts of the world. This growth has been
characterized by a dramatic reduction in the number of farm operations producing
livestock, and a big increase in the number of animals per farm unit. Finally, there has
been a geographic concentration of these intensive units, particularly in the Southeast
and mid-Atlantic states. Manure production in these intensive facilities simply
exceeds the capacity of nearby farmland to utilize it in plant production, resulting in a
major disposal issue and hence a threat to ground and surface waters as well as a
problem with local air pollution.

These structural changes in the industry led to the present CAFO rule that was issued
in December of 2002. This rule focused on the largest operations that represent the
greatest environmental threats. These units are required to implement comprehensive
nutrient management plans and to submit annual reports summarizing their
operations.

What are the environmental issues?

The manure from livestock operations produces a variety of potential pollutants
which can migrate to ground water, streams, rivers, and lakes. These pollutants
include nitrogen, phosphorus, sediments and organic matter, heavy metals, salts,
hormones, antibiotics, pesticides and pathogens (over 150 pathogens found in manure
are human health risks). Further, CAFO facilities release a variety of gases and
material into the atmosphere including particulates, methane, ammonia, hydrogen
sulfide, odor-causing compounds, and nitrogen oxides.

Of the water-polluting materials, which are covered in the CAFO rule, excess
nutrients can cause direct impacts on human water supply through excess nitrates,
impacts on agriculture through excess salts in irrigation waters, as well as
eutrophication of water bodies, anoxia and toxic algal blooms. These latter effects can
result in fundamental changes in the structure and functioning of aquatic ecosystems
including cascading effects that reduce water quality and species diversity.
Uncontrolled releases of animal wastes have resulted in massive fish mortality.

Pathogens in polluted waters are a health hazard, both directly and through the food
chain, for example, crops and shellfish. The potential human health impacts of
antibiotics and hormones in wastes have not been well identified but are of concern.

How were the environmental impacts quantified?

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Of all oftlie potential en\ironmeiilal impacts, the (WTO economic benefits analysis
focused to a I a rue extent on the nutrient runoff from land where manure has been
applied and quantifying the economic benefits that would accrue from the manure
management requirements of the (WTO rule To do so they utilized the Cil.l-.WIS
model ((iroundwater Loading I-fleets of Agricultural Management Systems) which
uses natural inputs of precipitation, radiation, temperature, and soil type and
management inputs of irrigation, crop type, tillage, fertilizer and pesticides The
outputs include nutrients, metals, pathogens, and sediments in surface runoff and
ground-water leachate This model was applied to model farms of different sizes,
animal types, and geographic regions l-roni this model the reductions in pollutant
loading of nutrients, metals, pathogens, and sediments were calculated for large- and
medium-sized (Al OS that would result from the application of the rule due to
nutrient management plans.

I low were the economic benefits \ allied '

Se\en categories of economic benefits were estimated water-based recreational use
(by far the largest category), reduced numbers of fish kills, increased shellfish
har\est. reduced ground water contamination, reduced contamination of animal water
supplies, and reduced eutrophication of estuaries Reductions in fish kills and animal
water supply contamination were \ allied using replacement cost. Increased shellfish
har\esls were \allied using estimated changes in consumer surplus W ater-based
recreation was \ allied using the Carson & Mitchell study described in Text IJox 2
The Aquaculture ITlluent (iuidelines abo\ e (iround water contamination was \ allied
using economic benefits transfer based on a set of stated preference studies There
was no national estimate of the economic benefits of reduced eutrophication of
estuaries, but there was a case study on one estuary focusing on recreational llshing
and using economic benefits transfer based on re\ealed preference random utility
models

A whole series of potential impacts were not included in the economic benefits
analysis that would relate to water quality impro\ ements of the rule, including human
health and ecological impacts of metals, antibiotics, hormones, salts, and other
pollutants, eutrophication of coastal and estuarine waters due to nitrogen deposition
from runoff, nutrients and ammonia in the air. reduced exposure to pathogens due to
recreational acli\ilies. and reduced pathogen contamination of drinking water
supplies These impacts were not monetized mainly because of a lack of models and
data to quantify the impacts and. in some cases, the lack of methods to perform the
monetization Other ecosystem impacts that were not considered include the potential
changes to aquatic ecosystem functioning that relate to their capacity to produce
goods of \alue to society

Te\( IJox 4: I lie Pmspecliu' l.conomic IkncTils of (ho ( k'.in AirAcl AiiioikIihoiHs

The first hospcctixc Benefit-Cost Analysis mandated by the ll^<> Clean Air Act
(CAA) Amendments included estimates of the economic benefits of protecting
ecosystems related to reductions in air pollutants to be expected from the amendments
(I S l-IW. ll)lW) The Agency included qualitali\e discussions of the following

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potential ecological effects of atmospheric pollutants haseel on a re\iew of 1 lie peer-
re\iewed literature (I S I-PA. ("liapler 7. and pp l>2-l>))

I iihlo 5: Table of Qu;ili(;i(i\c Discussions of Poicnlhil llcolo^iciil r.lTccls of Atmospheric
Polliiliinls Discussed in llic l-'irsl Prospcc(i\c licncfil Cosl An;il\sis (I'W))

Polliilsinl

Acute KITccls

l.onu-lorm KIToels

Acidic
deposition

Direct toxic elVects
to plant lea\ es and
aquatic organisms

Progressi\e deterioration of
soil quality

Chronic acidillcation of
surface waters

\ilrouen
deposition

Saturation of terrestrial
ecosystems with nitrogen
Progressixe eiirichmeiil of
coastal estuaries

Mercury,
dioxins

Direct toxic elVects
to animals

Persistence in
hiogeochemical cycles
Accumulation in the food
chain

Ozone

Direct toxic elVects
to plant lea\ es

Alterations of ecosystem
wide patterns of energy
llow and nutrient cvclinu

The Agency used two criteria to narrow the scope of work for <.|uanlillcalion of
impacts

•	The endpoint must he an identifiable ser\ ice llow

•	A defensible link must exist between changes in air pollution emissions
and the quality or quantity of the ecological ser\ice llow. and quanlilati\e
economic models must be a\ ailable to monetize these damages

The Agency pro\ ided estimates of three categories of economic benefits related to
ecosystems based on standard economic models and methods

•	I'conomic benefits to commercial agriculture associated with reductions in
ozone.

•	l-cononiic benefits to commercial forestry associated with reductions in
ozone.

•	l-cononiic benefits to recreational anglers in the Adirondacks lakes region
due to reductions in acidic deposition

l or agriculture, the Agency used crop yield loss functions from the National Crop
l.oss Assessment Network to estimate changes in yields. These yield elVects were
then ted into a model of national markets tor agricultural crops (A(jSl.M.) to estimate

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changes in consumers' and producers' surplus. The Agency did not quantify or
monetize effects on ornamental plantings. nurseries, or flower growers

Tor commercial forestry, the Pn 111-11 model was used to estimate the effects of
ele\ ated ambient ozone on timber growth The PnllT-lI model is a monthly time step
canopy to stand Ie\ el model of forest carbon and water balances based on maximum
net photosynthesis as a function of foliar nitrogen content The model relates ozone-
induced reductions in net photosynthesis to cumulali\e ozone uptake Analysis of
welfare effects used the I SI)A forest Ser\ ice Timber Assessment Market Model to
translate the increased tree urowth from a reduction in ozone to an increase in the
supply of har\ested timber and computed the changes in economic surplus
(consumers plus producer surplus) based on the associated price chanues liecause of
the lack of data and rele\ant ecological models, the Auency did not quantify or
monetize aesthetic effects, energy flows, nutrient cycles or species composition in
either commercial or non-commercial forests

for estimating the recreational economic benefits of reducing acid deposition in
Adirondacks lakes, the Agency used a published study of recreational angling choices
of households in New York. New Hampshire. Maine, and Vermont (Montgomery and
Needelman. Il^7) This was a random utility model of site choice Measured pi lof
lakes was used as an indicator of the I e\ el of ecological ser\ices from each lake The
literature on the economics of recreational angling shows that likelihood of success as
measured by numbers offish caught is a major determinant of demand for
recreational angling (see Phaneufand Smith 12< >( »51 and freeman | |W5| for re\iews)
To the extent that populations of target species are correlated with pi I le\els. pi I will
be a satisfactory proxy for llsh populations and angling success rates There was no
attempt to quantify other ecosystem ser\ ices of water bodies likely to be a Heeled by
acid deposition

Modeled reductions in acidification were used as an input to the Montgomery -
Needelman (ll^7) site choice model to simulate the effect of reduced acidification on
angler choice and angler welfare This simulation requires access to the data used to
estimate the model because the economic benefit measures to anglers depend on
indi\idual anglers' tra\el costs and site alternates

The Agency also presented an estimate of the economic benefits of reducing nitrogen
deposition in coastal estuaries along the east coast of the I S In order to estimate the
economic benefits of reduced nitrogen deposition in coastal estuaries, it would be
necessary to carry out the following steps

I. lislimale the changes in nitrogen deposition The Agency was able to do this

for the three estuaries co\ered in the Prospecthe Analysis
2 I se appropriate ecological models to estimate the changes in the populations
of species of concern to people These species include llsh and shellfish
species that are targets of commercial exploitation, llsh species that are targets
of recreational anglers, and perhaps other species that are of concern to people
such as birds and marine mammals Decreasing atmospheric deposition of
nitrogen was expected to reduce the deterioration of breeding grounds for
fisheries and reduce the habitat loss for aquatic and a\ ian biota It might be

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necessary not only lo estimate population changes for species that are resident
in and exploited within the estuaries hut also for species that use the estuaries
lor reproduction and shelter ofyoung or that are dependent on species from
these estuaries as a food source at some stage in their lilc cycle
3 I'Stimate people's willingness to pay lor increases in the ser\ ices pro\ ided In
these species There are models that can he used to do this for commercial
and recreational fisheries lint there is \ery little data on willingness to pay
for other types of ser\ ices such as hird watching and whale watching

The Agency was unahle to establish the necessary ecological linkages to quantify
these recreational and commercial fishery effects I lence it resorted to an a\oided
cost or replacement cost measure of economic benefits Reductions in nitrogen
deposition reaching l.ong Island Sound. Chesapeake Bay. and Tampa liay were
estimated The assumed a\ oided costs were the costs of achie\ ing ei|iii\ alenl
reductions in nitrogen reaching these water bodies through control of water
discharges of nitrogen from point sources in these watersheds As noted in Chapter 4
of this report. a\oided cost is a \alid measure of economic benefits only under certain
conditions, including a showing that the alternati\e whose costs are the basis of the
estimate would actually be undertaken in the absence of the en\ironmental policy
being e\aluated. that is. that the alternati\e's costs would actually be a\oided. Since
it was not possible to make this showing in the case of controlling nitrogen
deposition, the Agency chose not to include the a\ oided cost benefits in its primary
estimate of economic benefits, but only to show them as an illustrati\ e calculation

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6.2. VALUATION FOR SITE-SPECIFIC DECISIONS

6.2.1. Introduction

The Environmental Protection Agency makes many decisions at the local level,
including permits (air, water and waste); policies that influence the boundaries for
establishing permits (e.g., impaired water bodies designations); and administrative orders
related to environmental contamination. The social and ecological implications of such
decisions, like the decisions themselves, generally are local in nature, affecting towns,
townships and counties rather than entire states or regions. Therefore, these decision
processes need to rely on valuation approaches that also are local in nature and are robust
enough to adapt to a range of local stakeholder interests.

The U.S. EPA Science Advisory Board staff, with assistance from the Agency's
National Regional Science Council, surveyed the regional offices to assess their needs for
valuation information related to Agency regulatory programs. Seven of the eight responding
regions indicated that they need information to help value the protection of ecosystems in the
management and remediation of contaminated sites (U.S. EPA Science Advisory Board Staff
2004). The committee's goal is to help direct the Agency in building the capacity to satisfy
that need. Thus, in this section the committee focuses on the regulatory processes associated
with one set of local decisions, the remediation and redevelopment of historically
contaminated sites. That focus includes discussion of the Superfund program and its efforts
to assess the contributions to human well-being from ecosystem services related to site
remediation and redevelopment efforts (Davis, 2001; Wilson, 2005). The discussion that
follows is applicable to any remediation and redevelopment processes for contaminated
properties that contain the following basic and common elements:

a)	Site identification identification, selection, and prioritization of sites

b)	Site characterization - establish site condition

c)	Site assessment - evaluation of risks and impacts

d)	Selection of remedial and redevelopment approaches

e)	Performance assessment - clean up and redevelopment

f)	Public communication - assessment results; proposed actions and outcomes

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The goal in this section is to explore how the use of valuation methods can positively

influence individual steps in a remediation and redevelopment process and lead to a better
outcome. As appropriate, individual valuation approaches or methods relevant to specific
steps are identified and discussed. This section of the report aligns its analysis with a white
paper funded by EPA's Superfund Program (Wilson 2005) to evaluate the potential of
valuation for redevelopment of contaminated sites. The white paper provides an assessment
of the improvement in ecosystem service and implied ecological value from the remediation
and redevelopment of Superfund sites. Although the Wilson paper doesn't actually perform a
formal valuation for any individual redeveloped property, it does provide a useful starting
point for further exploration of the utility of valuation methods in the remediation and
redevelopment process. In preparation for his analysis, Wilson (2005) reviewed
approximately 40 superfund cases before selecting three case studies that represent urban
(Charles George Landfill), suburban (Avtex Fibers), and exurban (Leviathan Mine)
environments. The committee has chosen to analyze and rely on these same three cases, as
well as an additional urban example, the DuPage Landfill, because it provides a useful
counterpoint to the Charles George Landfill example. The DuPage example shows how an
early focus on ecosystem services can more completely identify potential ecosystem services
that can be targeted during the remediation and restoration phases. A brief overview of each
of these cases is provided in Text Box 6 through Text Box 9 below.

6.2.2. Opportunities for using valuation to inform remediation and redevelopment decision.

The Superfund process and its individual steps or stages are well defined (U.S. EPA

CERCLA Education Center, 2005). Superfund and related remediation processes are
focused on first defining a problem; then characterizing and assessing its potential and actual
human health and environmental impacts; and finally developing and executing a technical
strategy to alleviate or avoid those impacts. Since 1985 EPA's Brownfield Program (U.S.
Environmental Protection Agency, 2004) has integrated consideration of an upstream
redevelopment focus into the remediation process. The Agency built the Reuse Assessment
tool (Davis, 2001) to integrate a focus on land use into the Superfund process. Integrating
remediation and redevelopment makes evident the need to bring value concepts and
considerations to the beginning of the process and carry them through the individual steps or
stages of the process. Net Environmental Benefit Assessment (NEBA) (Efroymson et al.
2004, see Text Box 5) is a recent advance in thinking that provides a framework for using

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valuation tools to inform the comparison of alternative remedial strategies based on net

impacts on ecological services. Similar efforts are needed for other steps in the remediation

and development process.

TeM Box 5: V'l l'.n\ iroiiinoiiliil IkncTil Aii;il>sis

As described by IToymson el ill (2<)i)3) "\el en\ ironmental benefits are the gains in
en\ironmental ser\ices or oilier ecological properties attained by remediation or
ecological restoration, minus the en\ironmental injuries caused by those actions Net
en\ironmental benefit analysis (\l-li.\) is a methodology for com paring and ranking
the net en\ironmental benefit associated with multiple management alternati\es

A \I-IJA lor chemically contaminated sites typically in\ ol\ es the comparison of the
following management allernalk es: (I) lea\ ing contamination in place. (2)
physically, chemically, or biologically remediating the site through traditional means:
(3) impro\ ing ecological \ alue through onsile and oflsite restoration allernalk es llial
do not directly focus on remo\al of chemical contamination, or (4) a combination of
those allernalk es

\I-IJA in\ol\es aclkilies llial are common to remedial allernalk es analysis for state
regulations and the ("oniprehensk e I ji\ironmental Response. Compensation and
Liability Act. response actions under the Oil Pollution Act. compensatory restoration
actions under Natural Resource Damage Assessment, and proaclke land management
actions that do not occur in response to regulations i e . \aluing ecological ser\ices or
other ecological properties, assessing ad\erse impacts, and e\aluating restoration
options

I'igure 7. taken from l-lYoynison el al (2<)()3). depicts the high-le\el framework for
\I-IJ.V It includes a planning phase, characterization of reference state, net
en\ ironmental benefit analysis of allernalk es (including characterizations of
exposure of effects, including reco\ery). comparison of\l-IJ.\ results, and possible
characterization of additional allernalk es Dashed lines indicate optional processes,
circles indicate processes outside the \I-IJA I'ramework Only ecological aspects of
allernalk es are included in this framework The figure also depicts the incorporation
of cost considerations, the decision, and monitoring and efficacy assessment of the
preferred allernalk e. although these processes are external to \l-li.\ "

Since \I!15A is a framework, the resources, data inputs, and limitations are associated
with whate\er ecological models and \aluation tools are selected.

Currently. \I-IJA is being applied at a local scale, although the size of some
contaminated properties and their impacts can extend to the regional scale (i e impact
of releases from a contaminated site to a watershed) \ I! I i A should be highly
adaptable to different le\els of data, detail, scope, and complexity

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As noted, a generic process that encompasses remediation and redevelopment would
include a series of steps or discrete activities that involve ecological valuation. Figure 8
represents a generic remedial process in which opportunities to include valuation concepts
and assessment methods have been identified. As is clearly shown, early recognition of
future uses and ecosystem services that matter to people carries through to inform assessment
of the site and the ultimate selection of remedial actions and redevelopment options.
Optimally, expressing expected or actual contributions to human well-being will lead to more
effective communication with concerned publics. The following sections discuss
opportunities and utility of adapting valuation methods to this new merged process.

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Figure 8: Integration of Valuation information with traditional process to achieve improved

performance.

Value Inputs

Future use analysis
including ecological
and commercial uses

Align analytical
endpoints with future
use and ecosystem
services

Change to Current Practice

Consider opportunities to
integrate remedial and
redevelopment actions

Compare cost and
benefits of remedial
alternatives

Communicate
proposed actions

and expected '
outcomes to public

Document
benefits
delivered

Site
identification

J

Site

Characterization

I

Site
Assessment

I

Remedy
Selection

I



Remedial
Action defined





1 \



Site

Redevelopment

,





Performance
Assessment

Data collected could include
ecological condition and
community preference for
ecosystem services from
future use options

Assessment focuses on risks

relevant to planned future
uses and ecosystem services

Remedy optimized for
redevelopment and consideration
of ecological reuse

More cost-effective
redevelopment strategies

Greater public
understanding of benefits
and support for project

Lessons learned captured to
improve future project
execution

Valuation methodologies can be most useful for identifying how a site and the current
or potential ecosystem services matter to the surrounding community. Such methods should

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be focused on determining what contributions to human well-being have been derived, or can

be derived, from the site and how potential effects on ecological components diminish those

contributions. When the ecosystem services that matter to people are well-defined and when

the assessments of ecological production and risk can be coupled with these specific services,

then the outcome is likely to be a remediation and redevelopment plan that is targeted on

what really matters to the local community. A key recommendation, therefore, is that

consideration of ecosystem services and their contributions to human well-being and other

forms of value be considered from the earliest stages of addressing contaminated properties.

Even as early in the management process as site selection or prioritization, tools that
allow for comparison among sites for their potential to provide ecosystem services could be
informative. Assessment of the contribution of ecosystems or ecological protection to human
well-being should be considered in the design of any site characterization plan. While a
typical site characterization is focused on the aerial extent of chemicals and their range of
concentration in site media (e.g., ground and surface water, soil, and biological tissue), a plan
that also collects information to define and assess ecosystem service flows would better align
ecological risk and economic benefit assessments, as well as other kinds of assessments of
contributions to human well-being. Aligning risk assessments and assessments of
contributions to human well-being should be a critical objective for the Agency. Alignment
will help assure that the remedial actions will address the restoration of the contributions to
human well-being derived from any important ecosystem service flows that have been
diminished or disrupted. Aligning risk assessment endpoints with ecosystem services should
result in multiple benefits, such as: a) improved alignment with community goals; b)
improved ability to perform meaningful assessments of economic benefits and other
assessments of contributions to human well-being; c) improved ability to communicate
proposed actions; and d) improved ability to monitor and demonstrate performance.

The success of remediation and redevelopment of contaminated sites depends in great

part on the degree to which the ecosystem services and associated contributions to human

well-being important to the community are either protected or restored. If, as recommended,

values have been broadly explored and effectively integrated into the site assessment and

remedy selection processes, then measures of performance will be apparent. Ecological

measures of productivity or the aerial extent of conditions directly linked in an

understandable manner to valued ecosystem service flows will be useful in tracking the

performance of remediation and redevelopment processes. Advancing the Agency's

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capability to do performance evaluation both in real time and retrospectively will help the

Agency better justify its overall performance record in the remediation and redevelopment of

contaminated sites.

Finally, the remediation and redevelopment of a property encompasses more than just
the biological, chemical, and physical sciences and engineering principles that historically
have underpinned the remediation process. Effective communication with stakeholders
actively participating in the remedial and redevelopment process and with the general public
is a critical element in the success of the management process. Both of these audiences bring
values to the table when they evaluate proposed actions or the results of any action taken. A
strong alignment between the ecosystem services valued by these audiences and the expected
or actual outcomes will facilitate effective communication.

6.2.3. Recommendations and discussion of valuation through illustrative site-specific

examples

Chapter 2, Section 6 of this report included high-level recommendations. The
committee recommended that ecological values and contributions to human well-being
derived from ecosystem services be considered from the outset when framing any analytical
process to support Agency decisions and associated actions. The recommendations direct the
Agency to broaden its consideration of the types of ecological values and align them with
what matters most to the people involved in or affected by the decision.

In the following text, the general recommendations of Chapter 2 are applied to
valuation at the site-specific level. The committee illustrates these site-specific
recommendations with lessons gleaned from a series of Superfund examples in urban
(Charles George and DuPage Landfills), suburban (Avtex Fibers) and ex-urban (Leviathan
Mine) contexts. Text Box 6 and Text Box 7 provide background on the urban landfill cases.
Text Box 8 and Text Box 9 provide background on the suburban and ex-urban cases
respectively.

Text Boy 6: Charles George Landfill

From the late 1950s until 1967, the Charles-George Reclamation Trust Landfill,
located 1 mile southwest of Tyngsborough and 4 miles south of Nashua, N.H., was a
small municipal dump. A new owner expanded it to its present size of approximately
55 acres and accepted both household and industrial wastes from 1967 to 1976. The
facility had a license to accept hazardous waste from 1973 to 1976 and primarily
accepted drummed and bulk chemicals containing volatile organic compounds

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(VOCs) and toxic metal sludges Records show thai o\er l.ooo pounds of mercury
and approximately 2.5<)i) cubic yards of chemical wastes were landlllled The state
ordered closure of the site in llM3 That same year, the l-PA listed the site on the
\Pl.and the owner filed lor bankruptcy. Samples from wells ser\ ing nearby
Cannongate Condominiums and some nearby piix ale homes re\ealed VOCs and
hea\y metals in the groundwater Approximately 5<)i) people li\e within a mile of the
site in this residential rural area. 2. loo people li\e within 3 miles of the site The
nearest residents are located I'm feet away Benzene, lelrahydrofuran. arsenic. 1.4-
dioxane. and 2-butanone. among others, had been detected in the groundwater
Sediments ha\e been shown to contain low lex els of benzol a )pyrene People lace a
potential health threat by ingesting contaminated groundwater Mint Pond Marsh.

I'M ill Pond. Dunstable Brook, and nearby wetlands are threatened by contamination
migrating from the site

l-P.Vs in\ol\emenl at the site began with groundwater testing conducted by I'PA
contractor l-cology and I-n\ironment. Inc during llMI and ll->NZ The site was
proposed for the National Priorities List (\PI.) on October 23. Il->SI. and finali/cd on
the \PI. in September ll->S3 In September llM3 l-PA also allocated fluids for a
remo\al action at the site to replace the Department of I ji\ironmeiilal Ouality
I jigineering s temporary water line with another temporary but insulated water line
Other remo\al work included construction of a security fence along the northwestern
entrance to the landfill, regrading and placement of soil co\er o\er exposed refuse,
and installation of tweke gas xents A remedial in\estimation and feasibility study
(RI I"S) was also begun in September llM3 The basis for the remo\al action was
documented in the lirst Record of Decision (ROD) issued on December Zl). llM3.

IIPA Wch Silo llisloiv:

httn vosemite ena.uo\ rl nnl nad nsl'I'52l'a5c3 I laSI'5cSS525(-»adc""5"b(-'3 I AISDZS
01)71 ^1 )Z54S7S5Z5 m »l)i><>44^S2"( )nen I )ocument

Tex I Box n: Dul'sige ( oiinij l.;in(Hill

The 4o-acre tract of land that is now the lilackwell l.andllll was originally purchased
by the DuPage County forest Preser\e District (l-'PI)) in llMo and is centrally located
within the approximately I .Zoo-acre IJlackwell forest Preser\e The landfill was
designed to be constructed as a honeycomb of one-acre cells lined with clay
Approximately Z Z million cubic yards of wastes were deposited in the landfill
between llWo and N73 The principal contaminants of concern for this site are the
\ olatile organic compounds (VOCs) I .Z-dichloroetheiie. irichloroethene and
tetrachloroethene. detected in onsite groundwater at or slightly abo\e the maximum
contaminant lex el (MCI.), l.andllll leachate contained all kinds of VOCs and
senikolatiles including benzene, elhylheiizeiie toluene, and dichlorohenzeiie. and
metals such as lead, chromium, manganese, magnesium, and mercury VOCs and
agricultural pesticides ha\e also been detected in prkate wells down gradient of the
site but at low I ex els Some metals (manganese and iron) ha\ e been detected abo\ e
the MCl.s in down-gradient prkate wells Post-remediation, the site now consists
mainly of open space, containing woodlands, grasslands, wetlands, and lakes, used by
the public lor recreational purposes such as hiking, camping, boating, lishing. and

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horseback riding. There are no residences on the FPD property, and the nearby
population is less than 1,000 people. The landfill created Mt. Hoy, which is
approximately 150 feet above the original ground surface.

EPA Web Site History:

http://www.epa.uov/superflind/prourams/recvcle/impacts/pdfs/dupaue.pdf)

Tex I Box X: A\(ex l-'ihcrs Silo

The A\ lex Snpei 111ncl site consists of 44<) acres located on the bank of the
Shenandoah Ri\er within the municipal boundaries of front Royal. \ A The site is
bordered on the east by a military prep school (grades 5 -12). on the south by a
residential neighborhood, and on the west by the Shenandoah Ri\er I-Tom llM<> to
closure in NX1.). industrial plants on the site manufactured rayon and other synthetics.
Tons of manufacturing wastes and by-products accumulated on the site, inllltrated
into groundwater under the site, and escaped into the Shenandoah Ri\er The A\le\
fibers site was proposed lor inclusion on the National Priorities l.ist on October 15.
NX4. and the site was formally added to the list on June l<>. I^SO l-PA began
remo\al acli\ilicsal the site in Ito address \ arious threats to human health and
the en\ ironmeiil The cleanup and restoration plan called lor most remaining wastes
to be consolidated on site and secured with a protectee material where needed, and a
thick soil co\er and \egetation known as a cap

front Royal is located in close proximity to the Appalachian Trail, the Shenandoah
National Park and (ieorge W ashington National forest, making it a major tourist
center for the lilue Ridge Mountains IJiologically. the A\ lex site contains some
residual forested areas, open meadows, small wetland areas, and more than a mile and
a half of frontage along the Shenandoah Ri\er The proposed Master Plan for
rede\ elopment. created through a formal nuiiti stakeholder group process. di\ides the
site into three areas a) a 24o-;icre Ri\er Conser\ ancy Park along the Shenandoah
Ri\er combining ecological restoration and conser\ation of nati\e habitats, b) a 25-
acre Acti\e Recreation Park with boat landings, picnic shelters, and a de\eloped
recreation area including a \ i si tor center and soccer fields, and c) a Ko-acre Ixo-
IJusiness Park, featuring the refurbished historic former A\te\ administration
building Cleanup of the Axtex site is ongoing, and the rede\ elopment plan is being
acti\ely pursued by local go\ eminent agencies and pri\ ale industry groups

EPA Weh Site History:

Imp, www epa uo\ superl'und accomp success a\te\ htm

Stakeholders' "Avtcx l-'ihcrs Conservancy Park Master Plan"

littn www.a\lex fibers com Rcdc\ elonment aUexWI-IS a\te\-Mp html

Text Box *>: l.e\ hilhiin Mine Supcrl'iind Silo

In May of Ziioo. ihe | ;p \ added the l.e\ialhan Mine site in California to the National
Priority l.ist (NPI.) of Superfund sites The site is currently owned by the state, but
from ll->5l until llH-«2 the mine was owned and operated by the Anaconda Copper

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Mining Company (a subsidiary of ARCO) as an open pii sulfur mine The mine
properly is (oo acres located in a rural selling near the Ne\ada border. 24 miles
southeast of Lake Tahoe The physical disturbance from the mine itself is about 253
acres of the property plus an additional 21 acres of National forest Ser\ice land The
site is surrounded by national forest. In addition, it lies within the aboriginal territory
of the W ashoe Tribe and is close to se\eral different tribal areas

The mine has been releasing hazardous substances since the time that open pit mining
began in the ll)5<>s Releases occur through a number of pathways, including surface
water runoff, groundwater leaching, and o\erllow of e\aporation ponds In
particular, precipitation flowing through the open pit and o\erhurden and waste rock
piles creates acid mine drainage (AMD) in the form of sulfuric acid, which leaches
hea\y metals (such as arsenic, cadmium, copper, nickel, and zinc) from the ore
These releases are discharged into nearby l.e\iathan Creek and Aspen Creek, which
flow into the Last fork of the Carson Ri\er. Pollution abatement projects ha\e been
underway at the site since llM3 Despite these efforts, releases continue today

The releases of hazardous substances from the mine ha\e significantly impacted the
area s ecosystem and the ser\ ices it pro\ ides In the 1»s structural failures at the
mine that released high concentrations of AMD into streams resulted in two large lisli
kills, and the trout fishery downstream of the mine was decimated during this time
More recently, data ha\e documented ele\ated concentrations of hea\y metals in
surface water, sediments, groundwater, aquatic in\ertebrates. and fish in the
ecosystem near the site This suggests that hazardous substances ha\e been
transmitted from abiotic to biotic resources through the food chain, thereby affecting
many trophic le\els A recent assessment identifies se\en categories of resources
potentially impacted by the site surface water resources, sediments, groundwater
resources, aquatic biota, flood plain soils, riparian \egetation. and terrestrial wildlife.
The assessment identified ll\e types of ecosystem ser\ ices that might be pro\ided by
these resources aquatic biota (including the threatened l.ahontan cutthroat trout) and
supporting habitat, riparian \egetation. terrestrial wildlife (including the threatened
bald eagle), recreational uses (including fishing, hiking, and camping), and tribal uses
(including social, cultural, medicinal, recreational, and subsistence)

The process of determining compensatory damages and de\ eloping a response plan
for the site in\ol\es a number of different stages for which information about the
\ al Lie of these lost ser\ices would be a useful input. Lor example, in accordance with
Natural Resource Damage Assessment (NRI)A) regulation under the Comprehensi\e
Ln\ironmenlal Response. Compensation and Liability Act (CLRCI.A). the trustees
for the site conducted a pre-assessment screening to determine the damages or
injuries that may ha\e occurred at the site and whether a natural resource damage
assessment should be undertaken This requires a preliminary assessment of the
likelihood of significant ecological or other impacts from the contamination
(corresponding to Step 2 in figure 2 of this report) The decision was made at that
time (July ll^S) to mo\e forward with a Type li NRDA. which in principle is a
decision to mo\ e forward with an assessment of the \ al Lie of the ecosystem ser\ ices
that ha\e been lost as a result of the site contamination A Type 1} assessment
in\ol\es three phases a) injury determination to document whether ecological
damages ha\e occurred, b) quantification phase to quantify the injury and reduction

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in ser\ices (corresponding lo Slop 4 of figure 2). and c) damage determination phase
to calculate the monetary compensation that would he required (corresponding to
Step 5 of rigure 2) In the I .e\ iathan mine case, the Trustees proposed using resource
et|Lii\ alency analysis (RI I.\) based on a replacement cost estimate of the lost years of
natural resource ser\ ices to determine damages for all impacted ser\ ices other than
non-tribal recreational llshing Tor this latter ecosystem ser\ice. they proposed using
economic benefit transfer to estimate the \ alLie of lost llshing days finally, in the
decision by N\\ about whether to list the site on the MM. and the subsequent Record
of Decision selecting a final remedy lor the site, information about the \ al Lie of the
ecological impro\ enients from cleanup could play an important role, although these
decisions are often based primarily on human health considerations

IIPA Weh Silo History:

linn, www cna uo\ slineiI'lincl sites nnl narl5So htm

l.cvi;ilh;in Mine N;ilan"n2"rinal.pdf

6.2.3.1 At the beginning of the remediation and redevelopment process, define the ecosystem
services and values important to the community and key stakeholders related to the
site.

The urban examples of the Charles George landfill and DuPage County landfill show
the difference in outcome that can be produced by engaging with the community at an early
stage to focus on the ecosystem services of importance to them. Although there was no
evidence of formal valuation methods at the onset in either example, the focus on ecosystem
services and the inclusion of additional experts (i.e., forestry experts) led to a more positive
outcome for the DuPage County community.

At the Charles George landfill, ecological values or future uses were not considered at
the start. The human health risks at this site were so salient at the time they were discovered
that they were the focus of the subsequent decisions. When the landfill site was capped and
the water system from the city of Lowell, Massachusetts, was extended to the affected
community, the health and safety concerns were addressed. Although the Record of
Decision was published over 20 years ago, the potential for ecosystem services remains
untapped.

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By contrast, the remediation and redevelopment of the DuPage County landfill site,

now known as the Blackwell Forest Preserve, appears to have been motivated largely by the
need to address existence values (e.g., the presence of hawks and other rare birds) and
recreational values (e.g., hiking, bird watching, boating, camping, picnicking, sledding). The
remediation effort succeeded. Listed as a Superfund site in 1990, "a once dangerous area is
now a community treasure, where visitors picnic, hike, camp, and take boat rides on the
lake."

The urban examples show that even the most rudimentary dialogue about future use
can lead to an outcome with greater service to the community. At the DuPage landfill site,
even a qualitative focus on the utility of ecosystem services led to the recognition that in a
very flat landscape, even a 150-foot hill, if properly capped and planted, would be a welcome
refuge for people as well as wildlife. The DuPage Forestry District had a sense of the
ecological potential of the area, particularly for hawks, and a sense that, where hawks
abound, birders will come to watch them. In this case, the difference was not one of
methodology as much as conception. Once planners understand an area has ecological
potential, it may be fairly easy to utilize qualitative differences to show likely quantifiable
consequences

The Avtex Fibers case provides an example of the importance of engaging key
stakeholders. At the Avtex Fibers site, the public complained about offensive sights and
smells and contamination of drinking water wells. Over several decades, local government
and environmental protection agencies conducted tests, filed thousands of complaints, and
took various regulatory actions that ultimately resulted in the location's listing and
designation as a Superfund site. Once the site was listed and a management process
established, a clear effort was undertaken to engage stakeholders through a multi-stakeholder
process in the development of the Master Plan. Although there was some consideration of
ecosystem services, it is unclear whether there was any systematic means of assessing the
ecological services that people cared the most about.

For situations like the Avtex Fibers site, deliberative group processes involving

stakeholders and relevant experts (including historians) could provide an effective approach

to identify and document ecosystem service values of most concern to stakeholders. In

framing the dialogue with stakeholders, methods such as Ecosystem Benefits Indicators or

the Conservation Value Method might have helped EPA's site managers understand the

potential ecosystem service potential from future uses. Those methods could also provide

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inputs for further valuation using other methods described in Chapter 4 (e.g., economic

methods or decision science approaches methods).

Defining the ecosystem services that matter to people requires a carefully constructed
and systematically implemented program that integrates the use of multiple methods to fairly
and faithfully reflect the perspectives of multiple stakeholders. There is no simple recipe for
accomplishing this task and no simple algorithm for calculating values and summing them up
to make a decision.

The Leviathan Mine is a good example of how EPA is often faced with the need to
consider a complex array of competing interests. In this case the Agency is faced with a
clear dichotomy between the ecosystem services valued by the full-time resident population
of American Indians and the community of occasional recreational users. Recreational users
would gain from the cultural services associated with hiking, fishing, and camping.

However, the Washoe tribe that lives in the area year-round values the ecosystem as a
provisioning service for food as well as for its spiritual and cultural services.

The Leviathan Mine case study additionally highlights the need to consider the
existence or intrinsic values of the ecosystem. The ecosystem near the Leviathan mine site
provides a habitat for threatened species such as the Lahontan cutthroat trout and bald eagle,
which many individuals might value. In considering site restoration or remediation, or
measuring damages from contamination at the mine, the Agency could miss the primary
sources of value if it limited consideration to use value and did not consider these other
sources of value as well.

For the Leviathan example, information about the impacts of greatest concern to

affected individuals might be obtained in at least three ways. The first would be to gather

information from them about the relative importance of the various services through focus

groups, mental models, mediated modeling, deliberative processes, or anthropological or

ethnographic studies based on detailed interviews. The second approach would be to gather

some basic information that could be used to judge the importance of different services. This

might be of the type used to construct Ecosystem Benefit Indicators, such as: water use data

for the Washoe tribe and others in the vicinity of the site ( e.g., sources, quantities, purposes);

harvesting information for the Washoe (e.g., what percent of their harvesting of nuts, fish,

etc., comes from the area impacted by the site); recreational use data (number of people

visiting the area of the national forest impacted by the site for hiking, camping, fishing,

wildlife viewing); data on flooding potential and what is at risk in the vicinity of the site; data

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on spiritual/cultural land-use practices by the Washoe. It is unclear whether some of the

other data exist or would have to be collected. The third approach would be to review related

literature and previous studies to learn about impacts of concern in other similar contexts.

For example, previous social/psychological surveys (not specific to this site) or other

expressions of environmental preferences (e.g., outcomes of referenda, civil court jury

awards, etc.) might provide insight into what people are likely to care about in this context.

Similarly, previous contingent valuation studies of existence value might provide some (at

least partial) indication of the likely importance of impacts on species such as bald eagles

(e.g., if studies show that existence value is large). Likewise, previous studies of the value of

recreational fishing (e.g., from travel cost models) could be coupled with use data to provide

an initial indication of the importance of the impact on recreational fishing.

Analysis of the values of disparate users for a site is needed to identify the aspects of
the site contamination of greatest concern to people and the related ecosystem services. It
may be a significant challenge to identify and address the interests of different groups in
restoration and redevelopment. In the Leviathan Mine case, it is likely that this would have
to be considered both for tribal and non-tribal individuals, since the sources of value are
likely to be different for these two groups. .

6.2.3.2 Involve the mix of interdisciplinary experts appropriate for valuation at different sites.

Interactions among experts and the affected publics form a key component of any

hazardous site assessment, planning, and implementation program. Ideally, collaboration
among all relevant experts [physical, chemical, biological scientists (ecology, toxicology
etc.), and social scientists (economists, social psychologists, anthropologists, etc.)] and
communication with affected stakeholders begin very early in the planning stages of
remediation and redevelopment and remain throughout implementation and post-project
monitoring and evaluation. A key point for collaboration among expert disciplines is the
development of alternative management scenarios, particularly translating physical and
biological conditions and changes at the site into value-relevant outcomes that can be
communicated to stakeholders.

The Leviathan mine case provides examples of the need for collaboration among
disciplines to understand how the human population's values are affected. Because of the
unique cultural and spiritual values associated with ecosystem services, anthropologists could
play an important role in characterizing the value of the ecosystem services to the Washoe

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Tribe. Similarly, economists or others seeking to estimate existence value for an impacted

species (e.g., fish) would need to work closely with ecologists to determine the likely impact

of any change (or proposed project) on that species (e.g., effect on fish population) so that the

change could be valued.

6.2.3.3 Construct conceptual models that include ecosystem services.

Ecological assessments associated with the remediation and redevelopment of

contaminated property will be most meaningful for decision making if they incorporate
ecological production functions that link remediation and redevelopment actions to
ecosystem services. Historically, such assessments were not conducted at the four sites
chosen by the committee. The examples did, however, provide illustrations of how
assessments using ecological production functions could have influenced the site-specific
results in a positive manner.

While it is now standard practice to develop a conceptual model in performing
ecological risk assessments for contaminated site evaluations, EPA analyses of adverse
impact have generally not been linked to ecosystem services in ways that enabled alignment
of ecological risk assessments with economic benefits or other assessments of existing or
foregone ecosystem services. The primary focus of the Agency's remediation efforts is to
control anthropogenic sources of chemical, biological, and physical stress that could lead to
adverse impacts to human health or the environment. Developing a conceptual model that
incorporates the linkage between ecological endpoints and community-identified services can
help guide valuation of ecological protection, leading to practical information for site
remediation and redevelopment.

The Avtex Fiber case highlights what EPA could gain from developing the capacity
to use conceptual models that integrate ecological and social value attributes of a site. A
noteworthy feature of the Avtex Fiber process was the development of a Master Plan, which
provided evidence that some ecosystem services were considered but no evidence that
ecosystem services were broadly considered. For example, early concerns about
contamination of groundwater and discharge of toxic substances into the Shenandoah River
focused attention on water quality. Aquatic basins constructed to contain contaminants on
site were designed to restore important ecosystem services, including providing safe habitat
for waterfowl, runoff control, and water purification services. In this regard, the plan implied
- but failed to quantify or document - a rudimentary ecological production function.

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The development of a conceptual model that incorporated consideration of ecosystem

services would have systematically facilitated greater integration of ecosystem services into

remedial design and future uses. Recreational and aesthetic services were clearly important

considerations for many features of the plan, but no evidence suggests that a comprehensive

ecological model identifying ecosystem services guided redevelopment at the site. As a

consequence, it is unclear whether the particular pattern of restored forests and wetlands,

developed recreation areas, and industrial parks produced the best possible outcomes for

protecting ecosystems and ecosystems services. Different siting and design of the soccer

fields, for example, might have returned the same recreational value while achieving greater

ecosystem services in the form of wildlife habitat, water quality, or aesthetic values for

visitors, nearby residents or both. The declared ecological, "green" focus of the industrial

park as a component of the Master Plan implies that ecological concerns were important in

the selection of industrial tenants and in the siting and design of facilities, but no ecological

model for achieving this goal, or monitoring progress toward it, was presented. This

omission left open the prospect that future industrial, recreational, and tourist developments

and uses at the Avtex site might simply substitute one set of damages to ecosystems and

ecosystem services for another.

6.2.3.4 Adapt current ecological risk assessment practices to ecological production to predict
relevant ecosystem services

As discussed in Chapter 3 of this report, development of a conceptual model should
be followed with predictive analyses of effects of EPA's actions on ecological services. To
some degree, EPA's Ecological Risk Assessment Guidelines and Framework (U.S.
Environmental Protection Agency risk Assessment Forum 1992 and 1998) have endorsed the
concept that ecological risk assessments need to be built on a conceptual model linked to one
or more assessment end points. Expanding ecological risk assessments to include
assessments of the services that matter to people may present technical challenges, given that
current ecological risk assessments are often dominated by the available toxicological data
for a limited range of species and for toxic responses from individuals in those species. Such
data will rarely link well to the ecosystem services that matter to a particular site-specific
decision.

The Agency will need to develop its capacity to adapt and apply models that

incorporate ecological production functions for contaminated sites assessments. These

models are the real bridge between risk estimates and subsequent injury or damage

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projections and provide a major piece of the puzzle to quantify and value the impacts of

chemical exposures under different remedial and restoration alternatives.

EPA's assessments are important not only for EPA decisions related to site
remediation and development but also for decisions by other federal agencies. Although
other trustee agencies, such as the National Oceanic and Atmospheric Administration
(NOAA) and the U.S. Fish and Wildlife Service (USF&WS), are the regulatory leads for
Natural Resource Damage Assessment (NRDA), the ecological risk assessments and
conceptual models produced by EPA in the remediation process are often the basis for
damage assessment. The extrapolation from risk to injury to damages is often a controversial
aspect of the dialogue between the Agency, its trustee partners, and the parties responsible
for the damages. The estimate of risk and the estimates of uncertainties associated with
chemical exposure and toxic response introduces controversy because these data are often
used as a surrogate for injury to the environment. The related damage claim, an expression
of the restitution for lost or forgone use of ecosystem services, is likely to be challenged.
Predictive ecological production functions play a critical role in such decisions.

The Leviathan mine case illustrates how the concept of ecosystem services has been
used and can be used in damage assessment and restoration, as well as some of the issues
associated with delineating ecological services using ecological production functions to
predict impacts on them. If EPA could effectively conduct assessments that incorporate
ecological production functions to predict impacts on ecological services identified in
conceptual models, those assessments would enhance the ability of resource trustees to
appropriately assess injury, define restoration goals, and calculate damages.

For Natural Resource Damage Assessments, impact or injury occurs when some
standard (e.g., water quality or drinking water standards in the Leviathon Mine, for example)
is exceeded. Impact or injury also could occur when toxic substances are present in a
concentration or duration sufficient to cause a loss of services to the general public or a loss
of services unique to the Washoe Tribe. Thus, the concept of ecosystem services plays a key
role in defining or focusing categories of possible injuries to further evaluate.

Similarly, the concept of ecosystem services underlies the use of Habitat Equivalency

Analysis (HEA) or the related Resource Equivalency Analysis (REA) to determine

compensation for damages. In principle, application of HEA requires a determination of the

flow of ecosystem services that would have been provided by a given site had it not been

contaminated. This flow is then compared with the ecosystem services flow resulting from a

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restored site or a site providing equivalent services. Ideally, the value of the ecosystems

services under the two would be equal. In order to apply this concept, it is necessary to

delineate and value the service flows.

How can EPA estimate the impact of relevant ecosystem services? The Leviathan
Mine Natural Resource Damage Assessment Plan (NRDAP) gives detailed information on
concentrations of key pollutants (particularly heavy metals such as cadmium, zinc, copper,
nickel, and arsenic) in surface water samples, groundwater samples, sediment samples,
samples of fish tissues, and insect samples at various distances from the mine site. These
concentration levels can be compared to concentration levels at reference sites (since
historical information for the site itself is not available), toxicity data from the literature, and
existing regulatory standards (e.g., water quality criteria or drinking water standards) to
evaluate the potential for impact. Importantly, none of these approaches can be a direct
demonstration of injury, which can only be truly measured through field observation and
tests. EPA must rely on surrogates for estimating impact.

Once the impacts on water quality, sediments, etc., have been determined, ecological
production functions translate these impacts into predicted changes in the flows of services.
Estimations of the site's impact on the fish population in the nearby water body would need
to be considered to determine if recreational fishing is likely to be significantly impacted,.
Such an analysis requires estimation of the impacts of the changes in water quality,
streambed characteristics, bank sediments and riparian vegetation on fish population, both
directly and through impacts on the insects on which fish feed. If elevated levels of arsenic,
copper, zinc, or cadmium are known to exist in insects and fish tissue, EPA must be able to
use this information to predict an overall impact on the fish population.

EPA has already developed complex ecological risk assessment modeling tools (e.g.,

TRIM, EXAMS, and AQUATOX) to estimate the fate and effects of chemical stresses on the

environment. In some cases, EPA has even coupled such exposure-effects models with

ecological production models to estimate population level effects (citation?). In many cases,

an ecological model that links ecological processes at a site to ecosystem services of interest

to that site will not exist., although it might be possible to adapt models from the literature to

fit local conditions with site-specific field data, if the scale and ecological components of the

site are similar, using the criteria described in section 3.31 of this report for in selecting from

among existing models. In the absence of such a site-specific model, how should EPA

proceed in looking at the impact on ecosystem resources or services? At this stage, EPA

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might look to the scientific literature for guidance on how sensitive the insects and fish

species of concern are to these types of stressors. It could then ask expert ecologists to judge

the likely magnitude of the impacts in this specific case. This would be akin to an ecological

impact transfer, which is similar to the notion of economic benefits transfer. The Leviathan

Mine NRDAP suggests this approach. As for any issue involving transfer of information

related to valuation, scientists must take into account the differences between the reference

site and the contaminated site and define and communicate the assumptions and limitations

of transferring information.

In addition, the Leviathan Mine NRDAP suggests studying the fish population
downstream of the mine and comparing it to the population in a reference location, assuming
a realistic reference site can identified. More generally, it also suggests comparing riparian
vegetation, the composition of the benthic community, and wildlife populations near the
mine and at an acceptable reference site. Such a comparison can aid in framing the types of
damages resulting from the mining activity (which is most useful in an NRDA policy frame).
Since reference sites and exposed sites may differ for a number of reasons not related to the
contamination, such a comparison may not directly predict the injury and will not take into
consideration the impact of proposed remedial actions on ecosystem services. Decisions
about remediation and restoration require analysis of proposed actions and it may not be
reasonable to assume that remedial actions will be 100% effective in restoring the ecosystem
services to their original level (presumed to be the level at the reference site). Comparative
analyses using ecological production functions are needed and can be facilitated through
ongoing use of comparative tools such as Net Environmental Benefit Analysis (Efroymson
et. al., 2004).

6.2.3.5 Define ecosystem services carefully and develop a standard approach for cataloging

and accounting for ecosystem services for site remediation and redevelopment.

There is a need for accounting rules to recognize and avoid double-counting or under-

counting the contributions to human well-being from ecological service flows. Ecosystems
and their numerous components are linked in an intricate and complex network of biological,
chemical, and energy flows. By looking at isolated impacts to individual organisms or
components and their associated services, the potential arises for double counting or
undercounting contributions to human well-being generated by Agency actions addressing
contaminated sites.

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For example, the listing of services (aquatic biota and habitat, riparian vegetation,

terrestrial wildlife, recreational uses, and tribal uses) in the Leviathan Mine case does not

seem to be very useful for sorting out the different things to be valued. It fails to identify

mutually exclusive services and seems to present a high likelihood of double counting. It

also does not seem to adequately distinguish between inputs and outputs. The significance of

protecting habitat or riparian vegetation, for example, is not clearly addressed. Is it because

society cares about the populations it supports? Or is it because these populations are an

input into something else of value, such as recreation? Consider the example of insect

populations. If society cares about the insects for their own sake, then this should be included

as an existence or intrinsic value. If they are valued because they are a food source for fish,

and society cares about fish, then there is value in the change in fish brought about by the

change in insects. But in the latter case, they should not be valued separately. EPA should

view both clean water and insects as inputs into the production of more fish and value either

the inputs or the outputs. In order to determine how to measure value, it first must be known

why society values insects or fish.

Similarly, the listing of services by Wilson (2004) shown in Table 6, based on the

U.N. Ecosystem Millennium Assessment (2005) definitions of ecosystem services, is not

very useful for valuation purposes and could create confusion in valuation. For example, it is

unclear how or where the use of surface water or groundwater for drinking would fit in

Wilson's list. Is the service "Freshwater Regulation" intended to include drinking water or is

it intended as an input into aquatic and other habitat-related services?

Tsihle (>: r.c(is\sk'in Sen ice M;ilri\ lor l.e\ hillion Mine (I'min \\ ilson. 2004)

Ecosystem
Function

Ecosystem Service

Regulating

D isln rbancc Mode nil io n

•	Flood prevention from on-site evaporation
ponds

•	Regulation of surface water runoff and river
discharge during snowmelt and heavy rain
events



Freshwater Regulation

•	Restoration of groundwater discharge beneath
the pit and waste-ore piles

•	Non-hazardous surface water drainage into
Leviathan Creek. Bryant Creek and East Fork
River



Wildlife Habitat

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•	Nursery , feeding, and breeding ground for
indigenous fish species including the
threatened Lahontan cutthroat trout

•	Restoration of habitat and feeding habitat for
the threatened Bald Eagle

•	Maintenance of riparian vegetation habitat for
mammals, birds, amphibians, and insects

Supporting

Soil Formation

• Restoration of productive floodplain soils in
the Leviathan-Bryant Creek watershed and the
East Fork of the Carson River

Provisioning

Food and Raw Materials

•	Edible freshwater fish

•	Pine nut harvesting by Washoe tribe



Ornamental Resources

• Raw material for traditional Washoe Tribal
crafts

Cultural

Recreation and Amenity

•	Improved hiking and camping opportunities

•	Recreational fishing



Inspirational and historic

•	Washoe Tribal heritage site

•	Spiritual and ritual uses such as spiritual
bathing, and cleaning religious implements

Perhaps a better delineation of ecosystem services for use in ecological production
functions would involve, as discussed in section 3.3.2. of this report, the identification of
directly experienced, measurable, and spatially and temporally explicit measures of services.
Such a list of ecosystem services might consist of the following elements:

a)	Water used by Washoe Tribe members and others for washing and drinking

b)	Non-consumptive use values of wildlife (e.g., people like to view bald eagles
and other species)

c)	Harvesting (hunting, nuts, fish) by Washoe tribal members

d)	Cultural, spiritual and ceremonial value of land used by Washoe tribal
members

e)	Flood control (e.g., reduction in flooding from snowmelt or runoff)

f)	Recreational services (e.g., fishing, hiking, camping)

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Values that are broader and do not meet the principles described in section 3.3.2.

(e.g., "existence" or intrinsic values (broadly defined, based on moral or other principles)

from threatened and other species (e.g., cutthroat trout, bald eagles, and other impacted

species of concern); the value of the natural process leading to ecosystem outputs, beyond

the value of the outputs themselves (e.g., preference for natural processes over man-made

ones, or native species over introduced species) could be discussed qualitatively as

considerations to supplement the quantitative ecological production function analyses.

6.2.3.6 Expand the variety of methods the Agency uses to assess in monetary and non-
monetary terms the services lost or gained from current conditions or proposed
Agency action.

Chapter 4 of this report provides an overview of a broad range of methods that could
be explored for assessing ecosystem services lost or gained from current conditions or
proposed Agency action in monetary or non-monetary terms. Currently, without such
valuation of options, the typical comparison of remedial alternative strategies includes two
tests: a) whether a remediation action controls risk to an acceptable level; and if so, then b)
whether it is cost effective. Under this scheme, if a proposed remediation action is adequate
with regard to risk reduction, the least costly is the obvious choice. Such an approach
decouples remediation and development, leading to a delayed development process, possibly
off-mark from what matters to key stakeholders.

If remediation and redevelopment alternatives are to be compared based on an
analysis of their contributions to human well-being, a number of methods can be used. As
mentioned previously, NEBA (Text Box 5) offers a conceptual framework for comparing
remedial and redevelopment alternatives on a basis of contributions to human well-being,
whether monetized or non-monetized. For example, the contributions to human well being
associated with different remedial and redevelopment alternatives could be derived through
methodologies such as Habitat Equivalency Analysis (HEA) or Resource Equivalency
Analysis (REA) that report results in ecological units over time (e.g., discounted service
acres years). The cost of creation or replacement of those ecological units can be estimated
in monetary terms (i.e. replacement cost). This approach does not provide a direct measure
of the value of ecosystem services, but it does support a comparison of the services provided
under different options. Alternatively, impacts of alternatives could be compared purely in
ecological terms, such as through use of Biodiversity and Conservation Values approach or
energy-based approaches.

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Comparison of remediation and redevelopment alternatives using economic valuation

methods might include hedonic pricing studies to determine the economic impacts of the
identified cleanup and redevelopment options on adjacent residential property values. New
contingent valuation studies or studies of the value of recreational fishing (e.g., from travel
cost models) could be useful in capturing in monetary terms some of the values lost or gained
related to options being considered. Models might be used to compare expected gains to the
local economy across the feasible set of redevelopment scenarios. Monetary/economic
assessments and models might also be used to estimate the expected long-term contributions
to the local economy from industrial development versus recreation or tourism-focused use
options. For the Leviathon Mine, Ecosystem Benefit Indicators, as discussed above, might
also be used to evaluate the impacts of different mediation or redevelopment options.

If stakeholders are involved in testing remediation and redevelopment alternatives,
their preferences for or weighting of alternatives could be assessed directly through decision-
aiding processes and information about ecosystem services derived. This would allow non-
monetary methods such as biophysical ranking methods to be used as to compare changes in
biodiversity, habitat quality, energy flow, and other indicators of identified and accepted bio-
ecological goals, expressed in their own biophysical terms, across the cleanup and restoration
and redevelopment alternatives. Formal social-psychological surveys of potential
recreational users, visitors, and tourists could measure the relative preferences (importance,
acceptance) across the restoration/redevelopment plans under consideration from the
perspectives of these important groups. Parallel economic or monetary assessments, perhaps
using contingent valuation or travel cost methods or both, could extend and cross-validate
survey results. Decision-aiding methods could provide dollar-denominated value indices to
facilitate analyses of trade-offs with development costs and between recreation, tourism, and
industrial development emphases at a site.

6.2.3.7 Communicate information about ecosystem services in discussing options for
remediation and redevelopment of sites

The committee advises EPA to explicitly address issues regarding ecosystem services

in communications about site remediation and redevelopment. Because non-technical

audiences often find scientific information obscure, information about ecosystem services

might be communicated effectively through the use of visual communication techniques.

EPA might make effective use of perceptual representations (e.g., visualizations of

revegetation options as viewed from adjacent homes and prominent tourist and recreation

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sites and passageways) to improve stakeholders' understanding of the implications of the

various restoration and redevelopment alternatives under consideration. For example, the

restoration plan for the Avtex site included replanting and encouraging re-growth of three

different forest types on appropriate locations within the site. Accurate visualizations of the

reforestation projects, including their expected growth over time, would be very useful for

communicating the implications of alternative plans to stakeholders. Achieving and

effectively using such visualizations would first require interactions between forest ecologists

and visualization experts (such as some landscape architects). These interactions could lead

to the creation of accurate and realistic representations of how the different forests would

look from significant viewpoints at different stages of the restoration program for each

management alternative. Psychologists, communications experts, or other relevant social or

decision scientists might create appropriate vehicles and contexts for presenting the

visualizations to relevant audiences. Technical computer graphics expertise might also be

useful in this context. Further interdisciplinary collaboration would be required if the

visualizations were to be accompanied by information about expected wildlife or other

ecological effects associated with each visualized forest condition. While this example may

seem to be an intricate, exhaustive process, many contaminated properties are under

redevelopment for years (decades in the case of Superfund projects). With proportional

resource allocations, this level of effort is likely appropriate.

If valuation concepts and techniques are incorporated early and often throughout the
contaminated property redevelopment process, the Agency should be prepared to
communicate with interested publics more effectively. Managers will be able to
communicate the reasoning behind their selection of preferred options if analyses effectively
integrate consideration of ecosystem services and their derived contributions to human well-
being into the selection of the remedial and redevelopment actions,. Demonstrating to the
public that there has been a focus on ecosystem services that matter to them, and the ability
to communicate in terms of those contributions as they relate to proposed actions, should
lead to greater public understanding of options and acceptance of the proposed plan.

Projected contributions to human well-being should make the selection of performance
measures relatively straightforward. Communicating the progress and challenges of the
redevelopment process should be facilitated by using performance measures defined in terms
of contributions to well-being that the interested public understands and accepts as important.

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6.2.3.8 Create formal systems and processes to foster a information-sharing about ecological

valuations at different sites.

The committee recommends that EPA should actively pursue the broad and rapid
transfer of experience with integrating valuation concepts and techniques into the
redevelopment of contaminated sites. The Agency will build its capacity to utilize valuation
to inform its local decisions through a systematic exchange of information about site-specific
valuations. The lessons learned from these trial efforts, whether successes or failures, need to
be shared widely across the Agency with the regions, program offices, and the tool builders
in the research organizations. The Agency can catalogue and share such experiences in a
number of ways, such as reports, databases or BestNets (computer-based networks of users
sharing best practices). The Agency is in the best position to know how to build off existing
information exchange systems. Regardless how it is done, the information should be shared
broadly.

6.2.4. Summary of recommendations for valuation for site-specific decisions

The committee advises EPA to pursue opportunities for ecological valuation to

support decisions about site remediation and redevelopment. To effectively value the
protection of ecological systems and services in this context, the committee recommends that
EPA:

•	Define the ecosystem services and values important to the community and key
stakeholders related to the site at the beginning of the remediation and
redevelopment process,.

•	Involve the mix of interdisciplinary experts appropriate for valuation at
different sites.

•	Construct conceptual models that include ecosystem services.

•	Adapt current ecological risk assessment practices to ecological production to
predict relevant ecosystem services

•	Define ecosystem services carefully and develop a standard approach for
cataloging and accounting for ecosystem services for site remediation and
redevelopment.

•	Expand the variety of methods the Agency uses to assess in monetary and
non-monetary terms the services lost or gained from current conditions or
proposed Agency action.

•	Communicate information about ecosystem services in discussing options for
remediation and redevelopment of sites

•	Create formal systems and processes to foster a information-sharing about
ecological valuations at different sites.

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6.3. VALUATION IN REGIONAL PARTNERSHIPS

6.3.1. EPA Role in Regional-scale Value Assessment

Many important ecological processes take place at a landscape scale, making regional

analysis an appropriate scale at which to analyze the value of ecosystems and services. For
example, understanding habitat connectivity on landscapes, water and nutrient flows through
watersheds, or patterns of exposure and deposition from air pollution in an airshed pose
issues larger than a particular site. Rather, they are specific to a regional area and thus
require regional-scale analysis. Publicly available spatially explicit data on environmental,
economic, and social variables have increased dramatically in recent years. At the same
time, the ability to display data visually in maps and to analyze spatially explicit data using a
variety of analytical models and statistical methods has similarly expanded. The increase in
data and methods has opened up new frontiers for regional-scale analysis of ecosystems and
their services. An active EPA extramural program in ecological research is under way for
regional-scale analysis of ecosystems and services. As part of that program, EPA has funded
research relating to restoration of water infiltration in urbanizing watersheds in Madison,
Wisconsin.; restoration of multiple ecosystem functions for the Willamette River, Oregon;
decision support tools to meet human and ecological needs in rivers in New England; and the
provision of multiple services from agricultural landscapes in the upper Midwest. Region 4
has developed a tool for regional ecological assessment (discussed in Section 3.3.2). Other
regions have undertaken assessments of ecosystem services as well. Great potential exists,
largely untapped to date, to use this type of analysis to aid regional decision-making.

Municipal, county, regional, and state governments make many important decisions
affecting ecosystems and the provision of ecosystem services. Examples include land-use
planning and watershed management. Local and state governments rarely have the technical
capacity, or the necessary resources, to undertake regional-scale analyses of the value of
ecosystems or their services or to incorporate these values into their decision-making
processes.

Regional partnerships offer the potential for expanding local, state, and EPA capacity
to value ecosystems and their services. EPA regional offices have many opportunities to
collaborate at a regional scale with local and state governments, regional offices of other
federal agencies, environmental non-governmental organizations and private industry.
Through collaborating with local governments, other federal agencies, and the private sector,

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EPA can enhance its environmental protection activities by engaging important local

stakeholders, gaining access to regional expertise, and gaining access to decision-making on

important regional-scale environmental decisions. Local public and private partners can gain

from access to EPA technical expertise and resources. Such partnerships can expand the

knowledge base for decision making and improve the analysis of the value of ecosystems and

services.

Unlike national rule making, where analysis is often constrained by specific
mandates, the regional level enjoys great latitude to experiment with novel approaches to
valuing ecosystems and their services. Such experimentation may lead to improved methods
and practices of valuation with potential positive impacts well beyond the region that
pioneers the innovations. EPA, for example, can use regional-level partnerships as a
mechanism for testing and improving various valuation methods that might ultimately be
used at the national level. There is also a downside of not having legal or statutory
requirements for EPA to undertake valuation of ecosystems or services at the regional scale.
EPA regional offices with limited resources and a long list of mandated activities may have
little time or ability to undertake such activities with local partners. In addition, there may be
limited expertise in regional offices for undertaking at least some of the crucial steps that the
committee recommends in carrying out valuation of ecosystems or services. For example,
few regional offices have economists on staff who can work on valuation exercises. Partly
for these reasons, many of the potential advantages of regional partnerships for valuing
ecosystems or their services at a regional level have not been realized to date.

In analyzing regional opportunities for partnerships, this section explores several case
studies that illustrate some potential approaches to regional partnerships and regional-scale
analysis of ecosystems and services, including cases from Chicago; Portland, Oregon; and
the Southeast Region. Case studies illustrate several general lessons about regional-scale
analysis of the value of ecosystems and services and the potential usefulness of regional
partnerships.

6.3.2. Case Study: Chicago Wilderness

Chicago Wilderness is an alliance of more than 180 public and private organizations.

It represents a bottom-up organization that reflects the views of its member organizations to
protect the environment in and around Chicago. No single decision maker or agency controls
or guides Chicago Wilderness. It pursues objectives, as defined by its members, through

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consensus. The overall goal within Chicago Wilderness, as stated in Page 7 of the Executive

Summary of its Biodiversity Recovery Plan is "to protect the natural communities of the

Chicago region and to restore them to long-term viability, in order to enrich the quality of life

of its citizens and to contribute to the preservation of global biodiversity." Chicago

Wilderness pursues its goals by attempting to create a green infrastructure to support

biodiversity and to maintain ecosystems and services linked to quality of life in the Chicago

metropolitan area.

As a member of the Chicago Wilderness, EPA Region 5 provides technical and
financial assistance and facilitates the partnership. EPA expertise in Region 5, particularly in
natural sciences, has contributed to quantifying ecosystem services and understanding how
potential stresses affect ecosystems and the provision of services. Chicago Wilderness has
produced several reports, including a Biodiversity Recovery Plan and a green infrastructure
map for the region.43 The Chicago Wilderness Web site (http://www.chicagowilderness.org/)
contains a complete chronology and links to many relevant documents, including the
Biodiversity Recovery Plan.

Chicago Wilderness is interested in the valuation of ecosystems and services, but is
only beginning to explore the opportunities for valuation in its activities. Members of
Chicago Wilderness enjoy only limited technical expertise and practical experience in
valuing the protection of ecological systems and services. EPA Region 5 also has limited
capacity to undertake economic analysis of the value of ecosystem services. No specific
legal authority mandates valuation of ecosystems or services as part of the work of Chicago
Wilderness. Though not required, quantifying values associated with the conservation of
green space and biodiversity could be helpful for Chicago Wilderness in meeting its own
stated objectives and in communicating its analysis to other groups and the general public.
The possible uses of additional valuation tools identified by Chicago Wilderness members
include the following options:

•	To inform decisions on the establishment of green infrastructure, including priorities
for acquisition of land by, for example, forest preserve districts or soil conservation
districts;

•	To assess the value of preserving ground water and other ecosystem services related
to clean water;

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•	To assess the relative value of investing in different research projects to establish

priorities for funding decisions;

•	To assess the relative value of conventional versus alternative development efforts
and to demonstrate conditions in which development decisions that have positive
impacts on the environment might be in the financial interest of the developer;

•	To communicate effectively with residents of the Chicago region regarding the value
of green infrastructure and biodiversity and how these are related to quality of life for
area residents.

In sum, Chicago Wilderness, like many regional partnerships, would gain much from the
ability to analyze the value of ecosystems and services, but is constrained by lack of expertise
and resources in doing so.

6.3.2.1 An Example of How Valuation Could Support Regional Decision-Making: Open-
Space Preservation in the Chicago Metropolitan Area

Valuation of ecosystems and services is often most useful when done in the context of
specific decisions affecting the environment. The committee chose a specific decision
context, county open space referenda in the Chicago Metropolitan area, to explore how this
report's approach to valuation could be useful to support regional decisions.

Voters in four counties in northeastern Illinois have passed referenda authorizing
bonds for land purchase for open space preservation or watershed protection. In November
1997, voters in DuPage County passed an open space bond for $70 million. In November
1999, voters in Kane County and Will Counties passed bond issues of $70 million for open
space acquisition or improvement. In 2001, the voters in McHenry County passed a $68.5
million bond for watershed protection. While these multi-million dollar bond proposals have
put a substantial amount of money into efforts to preserve open space and ecological
processes in the region, they are insufficient to provide adequate protection for all
worthwhile open space or watershed protection projects. Given this shortfall, input about
what lands should be purchased, or what management actions should be undertaken to
maintain or restore natural communities would help to ensure that counties invest these funds
wisely.

This section of the report therefore looks at how valuation could help inform
conservation investments under the local county bonds. For this example, three types of

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values derived from protecting natural systems will be examined: a) species and ecological

systems conservation; b) water quality and quantity; and c) recreation and amenities. The

water quality and quantity discussion will focus on McHenry County because the bond issue

there related directly to watershed protection. The example follows the process outlined in

Chapter 2 of this report. The following sections describe: a) the process of stakeholder

involvement and input into defining values of ecosystems and services of interest; b)

predicting ecological impacts in terms of changes in ecosystem services; and c) using

methods to assess and characterize the values of ecosystems and services.

6.3.2.2 Process of Stakeholder Involvement Scientific and Technical Input and Public

Participation

The planning documents and activities of Chicago Wilderness reflect several of the
themes from Chapter 2 of this report, including interdisciplinary collaboration and broad
involvement. The Chicago Wilderness Biodiversity Recovery Plan (1999) discusses specific
roles for private property owners; local, state, and regional governments; intergovernmental
agencies; and federal agencies. The document also highlights the actions of EPA that affect
biodiversity and EPA's role in Chicago Wilderness.

Chicago Wilderness provides an excellent example of an organization that has made
extensive efforts to engage the local community in determining the most important features
of ecosystems and services in the region. Two of the great strengths of Chicago Wilderness
are the broad range of groups included and the commitment to open processes. This
inclusion allows the participants themselves to define the objectives, goals, and priorities of
the organization. The open and democratic process and the extensive efforts to include
multiple views and voices results in the group's goals and objectives being largely reflective
of what people in the region view as important to conserve in their region. Engaging local
communities is a vital first step in the process of valuing ecosystems and services.
Engagement helps to focus scarce agency resources on issues of prime local importance as
well as to promote partnership and dialogue.

The inclusive planning process endorsed by Chicago Wilderness includes developing
a common statement of purpose, setting up three working groups (steering, technical, and
advisory committees), and working through nine planning steps (from visioning,
development of inventories, and assessment of alternative actions, to adopting a plan).

Chicago Wilderness conducted workshops and meetings to define implementation

strategies and to prioritize among its long- and short-term goals, which focus on the

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restoration and conservation of biodiversity. For priority setting, several of the workshops

included non-monetary valuation exercises with qualitative rankings of importance. The

Biodiversity Recovery Plan also references other measures, such as polls and The Nature

Conservancy's global rarity index. In one 1996 poll, only two out of ten Americans had

heard of the term "biological diversity." Yet, when the concept was explained, 87% indicated

that "maintaining biodiversity was important to them" (Belden and Russonello 1996 as cited

in the Chicago Wilderness Biodiversity Recovery Plan, p. 117).

Chicago Wilderness also conducted eight workshops to assess the status and
conservation needs with regard to natural communities in the area: four species workshops
addressing birds, mammals, reptiles and amphibians, and invertebrates; and four consensus-
building workshops on natural communities addressing forest, savanna, prairie, and wetland.
The natural communities workshops developed overall relative rankings based on the amount
of area remaining, the amount protected, and the quality of remaining areas that incorporate
fragmentation and current management. The workshops also assessed relative biological
importance for community types, based on "species richness, numbers of endangered and
threatened species, levels of species conservation, and presence of important ecological
functions (such as the role of wetlands in improving water quality in adjacent open waters)"
(Biodiversity Recovery Plan Chapter 4, p. 41), and identified visions of what the areas should
look like in 50 years. The workshop participants judged the data as insufficient to allow
quantitative assessment of natural communities.

Two different groups of scientists and land managers identified a classification
scheme for aquatic communities based on physical characteristics. The groups assigned
recovery goals to streams (protection, restoration, rehabilitation, and enhancement) and
priorities to lakes (exceptional, important, restorable, and other, based on Garrison 1994-95).
Streams were assessed using the index of biotic integrity, species or features of concern, the
Macroinvertebrate Biotic Index, and abiotic indicators. The workshops also assessed threats
and stressors to streams, lakes, and near-shore waters of Lake Michigan.

Fostering public support through education and outreach is also an explicit goal of
Chicago Wilderness. The group emphasizes working with schools (including universities); it
also identifies individuals, agencies and organizations as targets for outreach and
involvement.

Chicago Wilderness' strengths in engaging local communities, however, also

highlight some of the difficulties involved in doing so. Different individuals and different

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member groups define value differently. Some groups care more about restoring pre-

settlement ecosystem conditions. Issues of open space and recreation are the primary

motivation for others. Others focus on maintaining water quality or conserving the region's

biodiversity. Because Chicago Wilderness is an organization based on consensus, the group

often cannot make choices involving trade-offs among worthwhile objectives. Protecting

biodiversity, protecting water quality, and providing open space and recreational

opportunities are all seen as good things. The choices become more difficult when getting

more of one goal implies getting less of another goal. The inability to make trade-offs

among objectives limits the ability of Chicago Wilderness to make policy recommendations

or have an influence on decision making. Valuation could help highlight which goals are of

greater importance and help decision makers navigate among difficult choices.

Another disadvantage of Chicago Wilderness' broad engagement of local
communities is the time consuming nature of community involvement processes. Chicago
Wilderness is not well placed to make rapid analyses or provide feedback on decisions that
occur over a short time period.

6.3.2.3 Predicting ecological impacts in terms of changes in ecosystem services:

Since Chicago Wilderness is committed to the value of protecting biodiversity, it is

interested in predicting impacts related to species conservation and conservation of
ecological systems at the landscape scale. Chicago Wilderness successfully applied a variant
of the Conservation Value Method to identify and prioritize conservation actions through
spatial representation and analysis of unique and threatened species and ecosystems. Use of
the method demonstrates how principles of conservation science can be used for planning
and how a transparent approach to mapping conservation goals can be useful in a regional
partnership. Chicago Wilderness' Biodiversity Recovery Plan describes in detail the
organization's conservation goals.

Water quality and quantity figure prominently in many ecological processes and in
the provision of many ecosystem services. Text Box 10 describes some effects on the
provision of ecosystem services that may result from the protection or restoration of
watersheds. In some instances, Chicago Wilderness and its member organization have
conducted prior studies making it possible to identify site-specific ecological characteristics
important to considerations of ecosystems and services.

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Text Box 10: Possible Ecological Impacts and Provision of Services from the Protection or
Restoration of Watersheds Based on the Work of Chicago Wilderness

Surface water

•	Availability—more water will be retained in the watershed because there is
less runoff from impervious surfaces

•	Periodicity of flows—changes in the hydrograph are mitigated because
precipitation will be captured in the soil and vegetation, and subsequently
released more slowly

•	Maintenance of minimum flows—there is a greater chance of maintaining
adequate minimum flows because of the dampening effects of intact
watersheds and continuation of subsurface flows

•	Flooding—flooding is reduced because of the retention capabilities of the
intact watershed

Subsurface water

•	Availability for domestic and industrial use—will be increased because
percolation and subsurface recharge will be enhance by natural soil surface
and vegetation

•	Maintenance of wetlands—those habitats that depend on the water table or
subsurface flow will be enhanced because natural percolation and recharge
processes will be maintained

Biological systems that depend upon water quantity

•	Special status species—increased persistence of those habitats that depend on
increased quantities of water in the watershed and containing protected
species

•	Specific habitats—increased water quantity and more uniform stream flows
will support regionally important ecological communities, e.g., in-stream
communities, bottomland forests, wetlands and wet prairies

Effect on water quality

•	Pollution dilution—increased flows will dilute concentrations of organic and
inorganic pollutants

•	Assimilation of biotic pollutants—increased stream flows will permit greater
opportunity for the assimilation of biological materials

To illustrate how Chicago Wilderness might characterize impacts on water quality
and quantity in McHenry County this report supposes that stakeholders and experts together
decided that the most important ecological services for comparing watersheds within the
county are: a) minimizing flooding; b) maintaining or increasing groundwater recharge; and
c) maintaining or increasing wetland communities. In reality, the most important ecological
services related to water would be determined by the stakeholder involvement and input
process discussed in Section 6.3.2.2.

To predict impacts related to flooding, Chicago Wilderness could make use of the

GIS database it collected, which includes layers depicting rivers, streams, wetlands, forest

lands, and floodplains. As a first approximation, historical records of flooding in McHenry

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County watersheds could be examined and watersheds with the greatest flooding could be

identified. The analysis could then evaluate the potential for restoring floodplain forests and

wetlands for mitigating flooding. To address whether groundwater resources would be

adequately maintained or increased by a development option, Chicago Wilderness could use

the maps of aquifers and soils in the GIS database that describe run-off and percolation rates

for each soil type. Watersheds could be compared in terms of potential for aquifer recharge.

The analysis could then consider the effects of alternative land use decisions on recharge

(Arnold and Friedel, 2000). To address whether wetland communities would be maintained

or increased, Chicago Wilderness could use topographic maps and GIS data on rivers,

streams, floodplains, forests, wetlands, and land cover to rank watersheds within McHenry

County in terms of potential wetlands minus current wetlands. The areas within watersheds

with the potential for expanding existing wetlands or restoring wetlands could be measured.

A number of GIS data files are available from McHenry County that could assist in
understanding how the protection of a given part of a watershed contributes to ecosystem
processes and services. What is often lacking, however, is a cause and effect relationship
that can be used to predict how provision of ecosystem services will change with alterations
in management or policy. It may be possible to transfer results from studies of ecological
services from other regions. For example, Guo et al. (2000) measured the water flow
regulation provided by various forest habitats in a Chinese watershed. If these relationships
are transferable, then estimates of the effect of a policy of restoring forest habitat on water
flow could be generated. Changes in water flow could then be used to predict impacts on
aquatic organisms and their production functions such as waterfowl, fisheries, and wildlife
viewing (Kremen, 2005).

The third set of values included in this example are recreational and amenity values.
Recreation covers a broad set of potential activities, from walking in the park to large game
hunting. Community input is required to establish what are important recreational activities
in the area. Access to parks and open space is a primary concern in many urban and
suburban communities. A study conducted in the Chicago Metropolitan Area found a
tradeoff between trying to locate open space close to people to provide access and locating
open space to conserve species (Ruliffson et al. 2003). Some recreational activities (e.g.,
fishing, hunting, bird watching) require input from ecological models, while others (e.g.,
walking in the park) may be more a function of location. Similar comments apply to amenity

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values where community input is important in determining what factors most contribute to

amenities in the eyes of the community.

Chicago Wilderness has done an admirable job of collecting spatially explicit
information relevant to land use, open space, recreation, biodiversity conservation, and water
quality and quantity issues. For this information to be relevant to decisions that affect
ecosystems, however, Chicago Wilderness needs cause-and-effect relationships that can
predict how policy choices would affect ecosystems and the provision of services. Chicago
Wilderness does not have the kind of information at its disposal that would allow it to
estimate ecological production functions. Chicago Wilderness can be quite effective in
providing descriptive information, particularly in the form of maps, but will be limited in its
ability to analyze alternative policies and make recommendations about which alternatives
are preferable. For example, it will be limited in providing analysis to a decision maker in
McHenry County concerning how to invest the $50 million approved by voters for watershed
protection in a way that will maximize the value of ecosystems and services, because it will
not be able to martial information about how particular actions affect systems and services
identified as important.

Gathering the necessary technical and scientific expertise to predict how policy
choices will affect ecosystems and the provision of services is a difficult task that introduces
another potential problem. The experts best placed to provide evidence may be tempted to
substitute their values on what is important for those of the stakeholders and community that
ideally set the objectives for the organization. For example, defining the levels at which
biodiversity targets can be considered as being met involves judgment. Different judgments
used in models may give rise to different sets of recommendations. Making sure that the
results of the analysis reflect the values of the community rather than the values of the
experts requires honest communication as well as commitment on the part of experts to carry
out the stated desires of the community faithfully.

6.3.2.4 Valuation of Changes in Ecosystems and Services in Monetary and Non-Monetary

Terms

When there are trade-offs among different services, (habitat protection versus
improvements in water quality, for example), information about the value of various aspects
of ecosystems and services is necessary to inform decision makers about what alternatives
are more beneficial for the community. This requires information about relative values that

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goes beyond understanding the ecological impacts of management and policy alternatives.

As noted in other parts of this report, the valuation of ecosystems and their services
can be conducted in numerous ways. This section begins with a discussion of the potential
contributions that valuation could make for Chicago Wilderness and briefly describes
possible valuation methods that could be applied for different types of ecosystem services.
This discussion goes well beyond what Chicago Wilderness has actually done in the
valuation realm. Chicago Wilderness has conducted very few valuation studies to date and
largely lacks the resources and the expertise to conduct valuations.

The overall goal of Chicago Wilderness is "to protect the natural communities of the
Chicago region and to restore them to long-term viability, in order to enrich the quality of life
of its citizens and to contribute to the preservation of global biodiversity." This goal was
derived with active input from member organizations and represents a consensus view of
their values. In some sense, the important valuation exercise for Chicago Wilderness was
carried out at the first stage in which Chicago Wilderness engaged the community and
gathered feedback on what the community felt was important. This process resulted in an
important statement about the values held by the collection of organizations that constitute
Chicago Wilderness.

Given this understanding and the clear statement of the overall goal of the
organization, formal valuation studies that try to quantify the monetary value of alternatives
may be of secondary importance. Of primary importance is to understand how various
potential strategies contribute to the protection and restoration of natural communities and
the ecosystem services they provide. The Conservation Value Methods could be used for
identification and prioritization of conservation actions that would contribute to this goal,
through spatial representation and analysis of biodiversity and conservation values. Chicago
Wilderness has devoted most of its attention to stakeholder involvement and biophysical
measures of the status of natural communities. It has devoted much less attention to
quantitative measures of value, monetary or otherwise.

With a clearly stated overall goal, such as "to protect the natural communities of the

Chicago region and to restore them to long-term viability," economic analysis may be largely

restricted to estimating the cost of various potential strategies to achieve that objective.

Information about how various potential strategies contribute to the protection and

restoration of natural communities along with information about the cost of these strategies is

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the main information necessary for cost-effectiveness analysis. Cost-effectiveness analysis

addresses the issue of how best to pursue an objective given a budget constraint. There is no

need to estimate the value of protecting natural communities or of ecosystem services.

Of course, things are rarely so clear. Even with a single overall goal, there are often
multiple dimensions and trade-offs among those dimensions that require the analyst to go
beyond cost-effectiveness analysis. For example, in protecting natural communities, there
may be trade-offs between protecting more of one type of natural community versus another.
When there are multiple natural communities of interest, or multiple ecosystem services of
interest, it becomes important to address questions of value, a practical matter when
investment of bond monies are at stake. Is it more valuable to allocate more resources to
restoring upland forest or wetlands? Is it more valuable to mitigate flood risk or improve
water quality? Such questions can only be addressed by comparing the relative value
attached to different natural communities or services.

Economic valuation of the protection of natural communities may be important for
Chicago Wilderness and the public at large for several reasons. First, when there are
multiple sources of value generated by protecting natural communities (e.g., species
conservation, water quality, flood control, recreational opportunities, aesthetics, etc.),
monetary valuation provides a way to establish the relative importance of various sources of
value. With prices or values attached to different ecosystem services, one can compare
alternatives based on the overall economic value generated. Second, some biological
concepts such as biodiversity are multi-faceted. How one makes trade-offs among different
facets of biodiversity conservation or among protection of different natural community types,
is ultimately the same question as how one makes trade-offs among multiple objectives.
Again, establishing prices on different components of biodiversity or on different natural
communities allows for analysis of trade-offs among components and an assessment of the
overall value of alternatives. Finally, monetary valuation may facilitate communication
about the importance of protecting and restoring natural communities in terms more readily
understood by the public.

Non-monetary valuation can also be used. If trade-offs among different natural
communities or among different services are needed, surveys containing attitude questions
may be helpful. It may be easier for people to answer attitude survey questions about
whether they think it more important to provide additional protection of forests versus

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wetlands, as compared to responding to questions about monetary valuation of forest

protection versus wetland protection.

Protecting natural communities may be done for reasons related to the provision of
ecosystem services or it may be done because people value intact natural communities (e.g.,
because they hold existence values or intrinsic values). The only methods currently accepted
by economists for estimating non-use values, such as the existence value of natural
communities or biodiversity, are stated preference methods: contingent valuation and
conjoint analysis. In trying to estimate of the value of protecting species and ecological
systems, Chicago Wilderness could survey respondents in the Chicago area using contingent
valuation or conjoint analysis. Alternatively, Chicago Wilderness could attempt to use an
economic benefits transfer approach by applying the results of relevant surveys done in other
locations. The advantage of obtaining a monetary value for the conservation of species and
ecological systems through contingent valuation or conjoint analysis is that it would allow
Chicago Wilderness to calculate a total economic value for alternative strategies. Without
using contingent valuation or conjoint analysis, Chicago Wilderness could not include non-
use values and would be able to estimate a partial economic value for each strategy.

Any effort to place a monetary value on non-use values through stated preference
methods raises the questions of whether monetary values are commensurate with the types of
values that Chicago residents attach to protecting natural communities. In discussing the
importance of protecting biodiversity, Chicago Wilderness emphasizes that a survey of
public attitudes regarding biodiversity involving Chicago focus groups found that
"responsibility to future generations and a belief that nature is God's creation were the two
most common reasons people cited for caring about conservation of biodiversity"
(Biodiversity Recovery Plan, p. 14). Contingent valuation of the bequest value of
biodiversity might be consistent with measuring "responsibility to future generations,"
although the respondents in the focus group were presumably thinking in moral rather than
monetary terms. Strong differences of opinion exist on whether it is appropriate to try to
capture such notions as "stewardship" or "moral values" in monetary terms using stated
preference methods.

Deliberative valuation exercises using citizen juries or other small focal groups might

be a particularly useful means of evaluating trade-offs among potential strategies to protect

natural communities in the Chicago Wilderness context. Under deliberative valuation,

experts would work with a small group of selected individuals in the Chicago area to

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determine comparative values for parcels of land through a guided process of reasoned

discourse. Deliberative valuation might enable participants to develop more thoughtful and

informed valuations, to better trade off among multiple factors, and to engage in a more

public-based consideration of values. Experts could use deliberative valuation either to try to

come up with monetary comparisons of the values of the alternative properties or with

weights that could be used to aggregate multiple layers of data.

Monetary values derived through deliberative valuations may differ considerably
from traditional private values, both because of the consent-based choice rules that
deliberative valuation employs and the explicitly public-regarded nature of the valuation
exercise. Recent analysis suggests that deliberative valuations may aggregate individual
values in a manner that systematically departs from the additive aggregation procedures of
standard cost-benefit analysis (Howarth & Wilson, 2006). Monetary values from deliberate
processes do not necessarily yield economic benefit measures.

As mentioned above, protecting natural communities may be done because people

value provision of ecosystem services (e.g., water quantity and water quality, recreation and

amenity values), as well as because they hold existence values or intrinsic values. Changes

in water quantity can be valued either because there is too much (flood control) or too little

water (water scarcity). One approach to measuring the value of flood control is to measure

avoided damages with reduction in probabilities of flooding. Several studies of the value of

preserving wetlands for flood control have been undertaken in Illinois including studies of

the Salt Creek Greenway (Illinois Department of Conservation, 1993; USACE, 1978) and the

value of regional floodwater storage from forest preserves in Cook County (Forest Preserve

District of Cook County Illinois, 1988). The later study found estimated flood control

benefits of $52,340 per acre from forest preserves. For water quality, an important

ecosystem service in many metropolitan areas is the provision of clean drinking water.

Protection of ecosystems may help reduce the fluctuation of water availability by storing

water during wet periods and gradually releasing it during dry periods. Ecosystems

protection may also be beneficial in providing relatively clean water for municipal supply.

There is also value of surface recharge of aquifers (NRC 1997). The value of providing

clean drinking water to the public is extremely high, far exceeding the costs of supplying it

whether by natural or human-engineered means. Because it is a question of how - not

whether - to supply clean drinking water, replacement cost (for example, the cost of building

a filtration system to replace lost water purification services provided by wetlands) can be

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used as a method to value the contribution of ecosystems to the provision of clean drinking

water.

A large literature in environmental economics exists on estimating the values of
various forms of recreational opportunities and amenities created by the natural environment.
Typical methods used by economists to estimate the monetary value of recreation and
amenities include hedonic property price analysis, travel cost, and stated preference. In
addition, a smaller literature uses evidence from referenda voting to infer values for open
space and other environmental amenities.

Applications of the hedonic property price model are a common method for
estimating the value of environmental amenities, especially in urban areas because of the
availability of large data sets on the value of residential property values. The hedonic
property price model has been applied to estimate the value of air quality improvements (e.g.,
Ridker and Smith 1967, Smith and Huang 1995), living close to urban parks (e.g., Kitchen
and Hendon 1967, Weicher and Zeibst 1973, Hammer et al. 1974), urban wetlands (Doss and
Taff 1996, Mahan et al. 2000), water resources (e.g., Leggett and Bockstael 2000), urban
forests (e.g., Tyrvainen and Miettinen 2000), and general environmental amenities (e.g.,
Smith 1978, Palmquist 1992). Given the large number of residential property sales in the
Chicago area and existing spatially explicit databases on many environmental attributes,
there is great potential for Chicago Wilderness to utilize such studies to estimate values of
various environmental amenities. This method has not been used by Chicago Wilderness to
date.

A large literature also exists on the value of recreation sites using the travel cost
method. With the large number of visitors to Lake Michigan beaches, forest preserves, and
parks in the Chicago metropolitan area, great potential exists for Chicago Wilderness to
apply travel cost to estimate the value of recreational activities. There have been several
applications of travel cost studies in urban areas (e.g., Binkley and Hannemann 1978,
Lockwood and Tracy 1995, Fleischer and Tsur 2003). To date, these methods have not been
applied by Chicago Wilderness.

Stated preference methods can also be used to estimate the value of recreational

opportunities and environmental amenities. One such study has been done for Chicago

Wilderness. Kosobud (1998) used a contingent valuation survey to estimate the willingness

to pay for the recovery or improvement of natural areas in the Chicago region. Kosobud

found an average willingness to pay for expanded natural areas of approximately $20 per

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household per year. Extrapolated over the number of households in the region, this would

generate about $50 million in benefits from expansion of natural areas in the region per year.

Finally, there is a small but growing literature that analyzes the results of voting
behavior in referenda involving environmental issues to estimate values. In particular,
studies have analyzed the value of open space using results of voting on open space referenda
(Kline and Wichelns 1994, Romero and Liserio 2002, Vossler et al. 2003, Vossler and
Kerkvliet 2003, Schlapfer and Hanley 2003, Schlapfer et al. 2004, Howell-Moroney 2004a,
2004b, Solecki et al. 2004, Kotchen and Powers 2006, Nelson et al. 2007). As noted, several
counties in the Chicago metropolitan area have passed referenda authorizing bonds to
purchase open space or for watershed protection. Though the number of referenda is
relatively small, making it difficult to generalize or make comprehensive statements about
values, analysis of the results of these referenda could provide insights into the values of
different segments of the public for various environmental amenities.

Application of valuation methods would generate quantitative estimates of the value
of the protection of ecosystems and the provision of various ecosystem services. This
information could be of great use to decision makers in evaluating alternative strategies to
protect natural communities. Valuation studies could also be quite useful in communicating
consequences of various alternatives to the public. Chicago Wilderness could usefully apply
a number of valuation methods for these purposes.

To date, however, Chicago Wilderness has initiated very little valuation research.
Despite some attempts to collect information about the value of protecting natural
communities and ecosystem services (e.g., Kosobud 1998), this effort has not been
comprehensive or systematic. This contrasts with the major efforts undertaken to garner
stakeholder involvement and input into setting the goals for the organization, and the large-
scale effort collecting technical and scientific knowledge to characterize the status of
ecosystems and species. In part, the lack of valuation activity is the result of the mix of
expertise of the individuals involved in Chicago Wilderness. In part, the lack of valuation
activity is the result of the choice made by the organization about the set of activities most
important to it (which is a different sort of revealed preference). Interest exists within
Chicago Wilderness to include economic and other social science approaches to study the
value of protecting natural communities, but the right mix of available expertise and
circumstances has been unavailable to make this a reality.

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6.3.3. Other Case Studies: Portland. Ore.; and the Southeast Region

6.3.3.1 Portland. Ore.. Assessment of the Value of Improved Watershed Management

The city of Portland, Oregon, facing potentially major expenses from meeting its

obligations under the Clean Water Act, Superfund, and the Endangered Species Act, decided
to invest in an analysis of ecosystem impacts and the value from ecosystem services that
would result from improved watershed management. By taking a systems approach and
considering the multiple economic benefits of actions, Portland officials hoped to find more
effective watershed management that would both save the city money and improve the
welfare of its citizens. Of primary interest were impacts on flood abatement, water quality,
aquatic species (salmon in particular), human health, air quality, and recreation. The City of
Portland's Watershed Management Program requested David Evans & Associates and
ECONorthwest to undertake the study, which they completed in June 2004 (David Evans &
Associates and ECONorthwest, 2004). Though not an example of a regional partnership
with EPA, the project provides one of the best current examples of the kind of landscape-
scale analysis of the value of ecosystems and services and exemplifies many of the
recommendations this report.

Portland city officials realized that they only understood a portion of the contributions
to well-being from improved watershed management. To be able to make intelligent
decisions about watershed management, these officials wished to have a more complete
accounting, which required applying methods that could quantify a range of ecosystem
values that are normally not quantified. The project aimed to expand the range of ecological
changes that are valued, focusing on those changes in ecosystems and their services that are
likely to be of greatest concern to people. From the beginning, the effort attempted to solicit
input from the public and important stakeholder groups about significant ecological impacts.
In addition to the value of direct flood-abatement impacts, the study monetized the economic
benefits of biodiversity maintenance, as represented by improvement of avian and salmon
habitat, air quality improvement, water quality improvement, by reduction of water
temperature, and "cultural services", which the study defined as including the creation of
recreational opportunities and the increase of property values.

In order to carry out the project, both biophysical and economics analyses were

commissioned. The biophysical analyses included studies of hydrology and flooding

potential, water quality, water temperature, habitat analysis for salmon and other aquatic

species, habitat analysis for birds and other terrestrial species along riparian buffers, and air

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quality impacts (ozone, sulfur dioxide, carbon monoxide, carbon, particulates). The

economic analyses included studies of the impact of ecosystem changes on property values

(including public infrastructure, residential and commercial property), the value of flood risk

reduction, the value of recreation, and the value of impacts on human health.

The project used an approach that closely resembles the ecological production
function approach advocated in this report The approach linked management changes, such
as flood project alternatives, to a range of ecological changes. These ecological changes
were analyzed for their effect on various ecosystem services. Finally, the economic analysis
attempted to value the changes in various ecosystem services. While the ecological and
economic analyses were largely conducted by separate teams, the project was designed to
provide a close linkage between ecological results and economic valuation.

Of particular note in this study was the emphasis on focusing the analysis to estimate
the change in values that would occur under various management alternatives. Rather than
provide a static description of current conditions, which is the predominant form of
information collected by Chicago Wilderness, the approach taken in Portland tried to
estimate cause-and-effect relationships that would allow the systematic appraisal of the set of
consequences of alternative policy or management decisions. This focus, along with a
systems approach capable of incorporating multiple economic benefits, made this an
effective vehicle to study the net economic benefits of alternative management options.

The Portland case provides a good example of the potential advantages of integrated
regional level analysis. The project undertook an integrated approach capable of analyzing
the impact of alternative management actions on ecological systems and the consequent
changes in the value of ecosystem services. Attempts were made to solicit input from the
public in the design of the project so that it captured the impacts of greatest interest to the
public. Results of the project were presented with a graphical interface that allowed
stakeholders to run scenarios and see the resulting impacts based on underlying biophysical
and economic models. The analysis effectively deployed existing methods and estimates, but
it did not attempt to develop or test new approaches or methods.

The project also illustrates some of the potential problems and limitations in

undertaking detailed quantitative landscape-scale analysis. Inevitably in this type of analysis

there are gaps in data and understanding. Gaps in understanding include how ecological

systems will be affected by changes in management actions, and how this will affect the

provision of ecosystem services and the consequent value of those services. For example,

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how will songbird populations change in response to changes in the amount and degree of

fragmentation of habitat? What is the value to residents of Portland of changes in songbird

populations? The study often had to use economic benefits transfer methods from cases

quite different from the Portland context to generate estimates of values.

Though the project was commissioned by the City of Portland and had minimal EPA
involvement, the project is a good example of the type of systematic and integrated approach
to valuing the protection of ecosystems and services advocated by this report. In particular,
the project aptly illustrates the sequence of steps, from input from stakeholders, through
characterizing change in ecosystem functions under various alternative policy and
management options, to valuation of services under alternatives. The project shows great
potential for this type of analysis to provide important and useful information to decision
makers.

6.3.3.2 Southeast Ecological Framework Project (EPA Region 4)

The Southeast Ecological Framework (SEF) project represents a regional GIS

approach for the identification of important ecological resources to conserve across the
southeastern United States. This region is one of the fastest growing regions in the country.
Even so, it still harbors a significant amount of globally important biodiversity and other
natural resources. The SEF is designed to meet EPA's goals of gathering and disseminating
information pertinent to the ecological condition of a region. The SEF project's goal is to
enhance regional planning across political jurisdictions and to help focus federal resources to
support state and local protection of ecologically important lands. The Planning and
Analysis Branch of EPA Region 4 and the University of Florida completed the work in
December of 2001.

This framework has been developed for the eight southeastern states in EPA Region 4
(Alabama, Florida, Georgia, Kentucky, Mississippi, North Carolina, South Carolina, and
Tennessee). This project has created a new regional map of priority natural areas and
connecting corridors, along with geographic information system (GIS) tools and spatial
datasets. The framework identified 43% of the land that should be protected and managed
for specific contributions to human well-being. Two additional applications of the SEF were
developed to demonstrate its utility for conservation planning at the sub-regional and local
scales. This approach is now being evaluated for utility in other regions and nationally.

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The SEF differs from the prior two case studies (Chicago Wilderness and City of

Portland) because it focuses on a broad regional analysis of eight states, rather than a single

metropolitan area or watersheds within a metropolitan area. The SEF also differs in that it

focuses almost exclusively on habitat conservation rather than a broad suite of ecosystem

services. The SEF did not undertake an extensive stakeholder involvement process to set its

objective; it started with the focus on habitat conservation. It also does not attempt to

combine economic analysis with ecological analysis to value the protection of ecosystems or

services in monetary terms. Discussion of values focuses on conservation value, which is the

ability to sustain species and ecological processes. In this regard, the SEF is a good tool to

carry out regional analysis of ecological components, particularly habitat conservation.

Because of its focus, the level of scientific knowledge underpinning the SEF is in general far

higher than in the other case studies. An important challenge facing regional analysis is

how to incorporate all of these essential elements: a rigorous ecological approach capable of

showing the range of ecological impacts from alternative policy and management decisions;

stakeholder involvement and input on what consequences are of greatest importance to them;

and rigorous evaluation of changes in value under alternative decisions, at a broad regional

scale like the eight-state Southeast region.

6.3.4. Summary and Recommendations

Regional-scale analysis has great potential to inform decision-makers and the public

about the value of protecting ecosystems and services. Recent increases in publicly available

spatially explicit data and a parallel expansion in the ability to display and analyze such data

make it feasible to undertake comprehensive regional-scale studies of the value of protecting

ecosystems and services. Municipal, county, regional and state governments make many

important decisions affecting ecosystems and the provision of ecosystem services at a

regional scale, but local and state governments rarely have the technical capacity or the

necessary resources to undertake regional-scale analyses of the value of ecosystems or

services. Regional-scale partnerships between EPA regional offices, local and state

governments, regional offices of other federal agencies, environmental non-governmental

organizations, and private industry could aid both EPA and regional partners. Such

partnerships offer great potential for improving the science and management for protecting

ecosystems and enhancing the provision of ecosystem services.

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At present, however, this potential is largely unrealized. The valuation of ecosystems

and services has not been a high priority for EPA regional offices largely because of tight

agency budgets and the lack of specific legal mandates or authority. To date, regional offices

have not undertaken the valuation of ecosystems and services at a regional scale in a

comprehensive or systematic fashion. As the case studies have shown, however, various

regional offices have pursued some innovative and promising directions despite limited

budgets and lack of specific legal mandates.

The committee sees great value in undertaking a comprehensive and systematic

approach to valuing ecosystems and services at a regional scale. Realizing the great potential

of regional-scale analyses, however, will require a significant increase in resources for

regional offices and, in some cases, a somewhat different mode of operation. To reach the

potential for regional-scale analysis of the value of ecosystems and services, the committee

makes the following set of recommendations:

•	EPA regional staff should be given adequate resources to develop expertise necessary
to undertake comprehensive and systematic studies of the value of protecting
ecosystems and services. Increased expertise is needed in several areas:

•	Economics and social science: Expertise is very limited at the regional level
to undertake economic or other social assessments of value. A pressing need
exists to increase expertise in this area among regional offices.

•	Stakeholder involvement processes.

•	Ecology: Regional staffs have greater expertise in ecology than in stakeholder
involvement, economics or other social sciences, but doing systematic
approaches to valuing ecosystem services will require additional ecological
staff. Of greatest utility would be ecologists with expertise in assessing
impacts on ecosystem services through ecological production functions to
evaluate alternative management options.

•	Integrated research teams: A systematic and comprehensive approach to valuing the
protection of ecosystems and services requires that ecologists and other natural
scientists work together with economists and other social scientists as an integrated
team. Regional-scale analysis teams should be formed to undertake valuation studies.
Teams composed of social scientists and natural scientists should participate from the

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beginning of the project to design and implement plans for stakeholder involvement,

ecological production functions, and valuation.

•	Community input and involvement: Gathering extensive stakeholder input is of great
importance to establish the set of ecological consequences of greatest importance to
the community at large. All regional-scale analyses of the value of ecosystems and
services need to include stakeholder involvement at an early stage to ensure that
subsequent ecological, economic, and social analyses are directed toward those
ecosystem components and services deemed of greatest importance by affected
communities. As the Chicago Wilderness example illustrates, different individuals
and different groups see ecosystems in different lights and have different objectives.
A good rule of thumb is to go bottom-up instead of top-down. In other words, it is
important to understand what various communities view as being valuable rather than
asserting what is valuable. An important question that should be addressed by EPA
regional offices is how to develop effective stakeholder involvement at broader
regional scales.

•	Misapplication of valuation: Some EPA staff have expressed a desire to be given a
value for an ecosystem component or service that they can then apply to their region
(e.g., a constant value per acre of wetland or wildlife habitat). Such short cuts to the
valuation process are typically uninformed by local social, economic, and ecological
conditions and often generate results that are not meaningful. This approach to
valuation should be avoided.

•	Information exchange: Regional staffs need to be able to learn effectively from
efforts to value the protection of ecosystems and services being undertaken by other
regional offices and extramural research. EPA regional offices should document
valuation efforts and share them with other regional offices, with EPA's National
Center for Environmental Economics, and with EPA's Office of Research and
Development. Each regional office should also be encouraged to publish their
studies.

•	Extramural research: Future calls for proposals for extramural research should
incorporate the research needs of regional offices for systematic valuation studies.
Doing so will maximize opportunities that future grant funding will be useful for
EPA's regional offices.

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1	• Regional partnerships: Regional staff should be encouraged to form partnerships

2	with local and state agencies or local groups where doing so advances the mission of

3	EPA directly or indirectly by promoting the ability of partner organizations to protect

4	environmental quality.

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7 SUMMARY OF MAJOR RECOMMENDATIONS AND

CONCLUSIONS

Among the most important benefits that EPA provides to society are the safeguarding
and enhancement of ecological systems and services. Any effort to value particular EPA
actions, whether in the form of a national regulation, local cleanup action, or regional
conservation partnership, that does not account as fully as possible for these ecological
benefits risks substantially understating the value of the actions. To date, however, most
EPA valuation efforts have provided only limited information about the ecological effects
and their value, focusing instead on the human health benefits of EPA's actions.

To ensure complete and accurate valuation of its actions, EPA must therefore
take a more comprehensive approach to assessing, valuing, and reporting on the
ecological effects of its actions. A more comprehensive approach can benefit EPA and the
nation in multiple ways. First, valuations that do a more complete and accurate job of
incorporating changes in ecological systems and services can help decision makers make
better, more informed decisions. Second, ecological valuation information can help EPA,
other governmental agencies, stakeholders, and the public as a whole to be more proactive in
taking actions that protect ecological systems and services. Third, a more comprehensive
approach can help EPA educate the public and stakeholders about ecological systems and
services and their importance to society. Finally, by providing for greater public
involvement in valuation efforts and communicating more clearly about the methods used, a
more comprehensive approach can build trust in Agency science and decision-making.

In valuing the ecological effects of an action or proposed action, EPA should
identify all of the affected services that are important to the public and then provide as
much information as practical about both the impacts of the action or proposed action
on these services and the value of these impacts. EPA in the past has sometimes worked in
the reverse direction - first identifying those services that it believed it could monetarily
value or that it has valued in previous assessments, and then focusing its valuation efforts on
these services. Such an approach risks ignoring those services that are of greatest societal
importance and under-representing the value of EPA's action to decision makers and the
public. Even where full monetary valuation of a specific service is impossible, EPA often
can provide other information of use to decision makers such as the likely biophysical

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changes in the services and the relative importance of the services to the public. EPA, in

short, should start any valuation effort by deciding what it should value, not what it can

most easily value.

The critical first step in ensuring that EPA's valuation efforts address all
relevant ecological effects is to begin all valuation assessments by developing a
conceptual model of the potential ecological changes. This model, which should be
constructed at a general level, can serve as a roadmap to guide both the identification of
relevant ecological effects and the development of more specific ecological analyses.
The model should include information and linkages for all relevant levels of the
ecological system, including information about both the important underlying ecology
and the linkages between the ecological outputs and the ecosystem services of
importance to society. Peer review of the conceptual model may be helpful in ensuring that
the model is sufficiently comprehensive to serve as the foundation for the valuation
assessment.

Guided by the conceptual model, valuation assessments should involve four key
steps. First, experts must predict the effect of policy-induced changes on basic
characteristics of the relevant ecosystems, using ecological models that are scaled and
parameterized to the ecosystems. Second, EPA must identify the ecosystem services
that are of public importance. These are the services that should be the focus of the
valuation assessment. Third, experts must map the predicted ecological changes to
changes in these ecosystem services. Finally, experts must quantify or characterize the
value of the changes in the ecological systems and services to the extent possible.

This multi-step process requires greater collaboration among ecologists,
economists, and other experts than has historically been found at EPA - as well as
greater participation by the public. In developing the initial conceptual model of the
potential ecological changes, for example, EPA should involve experts in both relevant
biophysical aspects of the modeling and social scientists. Through the use of mediated
modeling and similar techniques, EPA should also involve the public, and incorporate public
views and understandings, in the development of the conceptual model. The identification
and development of effective measures of ecosystem services also requires input from
ecologists, who know what biophysical changes can be measured; social scientists, who
know what can be valued; and the public, who can help in the identification of those services
of social importance.

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One of the principal questions that EPA faces in the first step of predicting ecological

effects is which models to employ. Because models are continuously being modified,

EPA should develop a set of clear criteria for the selection of appropriate models in

each valuation exercise rather than dictating the use of particular models. These criteria

should draw on prior work by EPA, the National Research Council, and other experts to

identify relevant factors to consider in choosing among alternative models.

Both the public and experts, as noted above, have a significant role to play in the

second step - determining which ecological services are important to the public. The

relevant question is what is of importance to the public, not to experts. Public surveys,

individual narratives, mental model research, focus groups, content analysis of public

comments, and similar approaches can help identify relevant public attitudes. The public,

however, often may not fully understand the nature or potential importance of

particular effects. Experts therefore have an important role to play in helping to

educate the public about the science of ecological systems and services, the role that

particular services play in advancing societal interests and goals, and the likely impact

of particular changes in those services on the public.

One of the critical gaps today in many valuation assessments comes in the third

step - identifying how the biophysical effects of an action on an ecosystem will in turn

impact the ecosystem services of importance to the public. Even where ecologists are able

to assess the likely ecosystem effects, ecological production functions often do not exist for

determining the quantitative relationship between changes in an ecosystem and changes in

the services that the ecosystem supports. A major priority in EPA's research therefore

should be the development of ecological production functions that can be effectively

applied in valuation assessments to predict changes in ecosystem services based on

changes in the underlying ecosystem. A number of research groups are currently working

to develop a first generation of models for measuring and mapping ecosystem services and

changes to those services under various scenarios. The committee believes that EPA can

significantly contribute to defining what types of ecological production functions and

models would be useful to its work and then supporting the further development of

these production functions and models.

As discussed in Chapter 4and Appendix B of this report, multiple methods exist not

only for identifying potentially important services but also for the final step in the assessment

process - valuing changes in ecological systems and services. EPA has historically focused

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on using economic methods to assess the value of such changes, and the Office of

Management and Budget requires that economic methods be used, where possible, in

valuations for Regulatory Impact Assessments of major rules and regulations. Economic

methods have the advantage in cost-benefit assessments of measuring the value of all costs

and benefits in a common metric that permits ready comparison of the costs and benefits of

particular actions. Economic methods also rest on a well-developed and consistent

theoretical framework, and significant economic data has been collected that may be of use

in developing monetary valuations in specific cases.

A number of additional methods identified in Chapter 4 also exist for assessing the
value of changes in ecological systems and services and have been used far less in actual
valuation efforts by EPA. Some of these methods (such as psycho-social survey of public
preferences) are grounded in substantial research and experience and are usable by EPA
today. Other methods (such as citizen juries) need additional research and development
before EPA considers using them as part of formal value assessment processes with
significant legal or regulatory consequences. The detailed descriptions of the various
methods contained in Chapter 4 provide guidance on the committee's views of the
strengths and weaknesses of each method, which methods are currently ready for EPA
use, and the research needed to strengthen and improve each method.

The committee believes that EPA can and should, where and to the extent

permitted by law, make greater and more sophisticated use of those methods that have

already been validated by substantial research and experience (including the survey

techniques discussed in Appendix C). Such methods can serve a variety of roles. Where

current economic methods cannot provide an accurate assessment of the economic value of a

particular change in ecological systems and services, psycho-social surveys and other proven

methods may provide decision makers with important and useful information on the value

that the public attaches to the actions that they are considering. In some situations (e.g.,

where decision makers are attempting to maximize a particular end such as biodiversity),

these methods may provide information of more direct relevance to decision makers or the

public in the context of the particular action or actions being contemplated. Providing

multiple measures of value also can be important in many settings. Because different

methods measure different aspects or concepts of value, the use of multiple methods can

provide decision makers with a more comprehensive and robust understanding of the value

of pursuing a particular action and thereby help them to make more informed decisions. In

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some cases, multiple measures of value may reinforce the societal importance of a specific

action; in other cases, the multiple measures may push in different directions, requiring

decision makers to weigh or balance the various types of value reflected by the measures. In

all cases, however, multiple measures will increase the information available to decision

makers.

EPA also should be proactive in identifying potential opportunities for testing,
using, and further developing new methods of valuation such as citizen juries or
ecosystem benefit indicators. Regional and local decisions, in particular, may present
settings where new methods can be appropriately tested and refined. In these settings,
legal mandates may not constrain the specific valuation methods that can be used, and the
decision making setting might be particularly suited to a new method. By seeking out
opportunities to use and test new methods, EPA can advance the understanding of these
methods and ultimately expand the set of established methods that it has available to use in
all settings. EPA also can help advance new methods by developing an extramural grant
program focused explicitly on this task.

In choosing which valuation methods to use in any particular setting, EPA
should recognize and take into account that different methods rest on different
assumptions and concepts of value. The economic methods that EPA has traditionally
used, for example, assume that the key values of importance to decision makers are the
monetary values that individual members of the public attach to particular ecosystem services
based on their role as consumers of such services. Several other methods of measuring
public value (e.g., measurements based on the results of initiative or referenda, and citizen
juries), by contrast, assume that members of the public attach different values when placed in
the role of citizen rather than the role of consumer. Various deliberative and assisted
methods assume that many people do not have well formed monetary values for ecosystem
services and that accurate valuation requires experts to actively assist people in constructing
and determining the value. EPA should be conscious of the different concepts of value
underlying various valuation methods and choose methods for particular assessments
based in part on which concepts of value are important or relevant to decision makers
in that context. As noted earlier, decision makers may often benefit from the
development of multiple measures of value.

In assessing and reporting value, EPA should also be as transparent and explicit

as possible as to what methods it has used, why it chose the methods that it has used, the

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assumptions underlying the methods, and the limits of the methods. One goal should be

to provide decision makers and the public with information about the assessment, the

choices underlying it, and the limitations to it that they need in effectively evaluating

the underlying action. A second goal should be to help decision makers and the public

understand how EPA derived the values embodied in the assessment report. From the

perspective of decision makers and the public, valuation assessments today can frequently be

black boxes that yield estimates of benefits and costs but little insight into the makeup of the

underlying estimates. Providing more information about the assessment methods can again

help decision makers and the public understand the relevance and credibility of the valuation

assessment.

Full quantitative valuation, whether in monetary or other metrics, of particular
changes in ecological systems and services of importance to the public is not always
possible. Where a full quantitative valuation is not possible, EPA should provide
decision makers and the public with as much biophysical information as possible about
the change and with other available information that can help decision makers evaluate
relevant actions and tradeoffs. Not surprisingly, OMB's Circular A-4 calls for exactly this
type of information when fully monetized valuation is impossible in a regulatory impact
assessment.

In these settings, EPA should pursue a hierarchy of information. Where a full
quantitative valuation is impossible, EPA should attempt to provide whatever
quantitative and qualitative information is available regarding the value that the public
attaches to the estimated ecological changes (e.g., information regarding the general
importance that the public attaches to biodiversity). Where no valuation information is
available, EPA should try to provide information on estimated change in ecosystem
services and the reasons that the services are of importance to the public. Where even
this information is not available, EPA should provide information on estimated changes
in the underlying ecosystem (e.g., in functional groupings of organisms) and how those
changes may affect connected ecosystem services. In this regard, EPA should develop
key ecological indicators that can be used in multiple contexts to characterize likely
changes in ecosystem services of public importance.

One of the most important but difficult issues involved in many major valuation

assessments, no matter what valuation method is used, is benefit transfer. Particularly in the

case of national assessments, information may not be available to directly value changes in

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ecological systems and services at all relevant locations. To assess value, experts therefore

often seek to use the results of benefit studies done in one location to estimate the benefits of

similar ecological changes in other locations and settings. In conducting such benefit

transfers, EPA must be aware of the limitations and risks of such transfers.

Inappropriate benefit transfers are often a very weak link in valuation studies.

Resource and information limits often make benefit transfer imperative. The

committee's review of various uses of benefit transfer, however, indicates that the drive

for numerical valuations may present the temptation to push benefit transfers farther

than they legitimately can be pushed (e.g., by using valuation estimates that are too old

to be reliable). EPA should consider both developing criteria and guidance for the use

of benefit transfer and establishing procedures (e.g., expert and in-house reviews) for

assessing and determining whether a benefit transfer is appropriate in a particular

situation. EPA should also support research that is likely to lead to a larger set of value

estimates of likely use in future benefit assessments.

Similar issues arise in the transfer of ecological information from one context to
another. Such extrapolation of ecological information requires caution, and agency
experts should carefully evaluate each proposed transfer to determine its
appropriateness. Transfers of ecological information may be more acceptable in some
contexts than others. Information is more likely to be transferable, for example, where there
is significant similarity between contexts or where the information is aggregate.

Valuation exercises are inevitably subject to a variety of uncertainties, and
valuation reports should include assessments of that uncertainty. Such assessments are
essential if decision makers are to make informed evaluations of proposed policies and
alternative policy approaches. Assessments of uncertainty also can help EPA to develop
research priorities for the improvement of valuation methods. Uncertainty assessments
should at a minimum report ranges of values and statistical information about the
nature of uncertainty for which data exist. Where possible, uncertainty assessments
should use formal quantitative methods, such as the Monte Carlo approach, which
provide a more reliable and rich characterization of the uncertainty.

How ecological benefits are communicated is also important, and EPA should

focus more attention on communication issues. One area where the Agency can

immediately improve its communication is in the characterizing of non-monetized

ecological effects in regulatory impact assessments. EPA, for example, should label total

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monetized benefits as "Total Monetized Benefits" rather than "Total Benefits" to clarify that

they do not include all ecological effects. EPA, moreover, should report the non-monetized

ecological benefits explicitly and, where possible, in units that are biologically appropriate

and socially relevant.

The resources needed to complete an accurate assessment of the value of changes
in ecological systems and services should not be understated. The committee
encourages EPA to develop a complete vision of what they would like to be able to
achieve in valuation assessments and then to seek the resources needed to do so.
Regional offices, in particular, need additional resources with which to produce
ecological benefit assessments of importance to local and regional decision making
either within or outside the Agency.

EPA can maximize the use of the experts and the resources that it currently
enjoys by providing for increased and improved information sharing about valuation
methods and the results of prior valuation studies and assessments. A number of EPA
regions are experimenting with valuation efforts in different settings, and the Agency as a
whole would benefit by creating a mechanism for widely sharing the lessons of these efforts.
Data compiled in one assessment process, moreover, may prove of value in another setting.
EPA therefore could also benefit by creating a mechanism for identifying and sharing
data - not only within the Agency but also from sources outside EPA. As part of this
effort, EPA should establish links with various efforts to collect relevant data, including the
NEON planning process and the NSF LTER program.

In conclusion, the committee recommends that EPA "think big" in valuing the
changes in ecological systems and services that flow from its actions. Too often in the
past, benefit assessments have under-accounted for such ecological changes. In order to
ensure that decision makers fully appreciate the benefits of EPA's actions, the Agency
must try to understand and assess such ecological changes as completely as possible and
in terms that matter to the public. This will require the use of a broader set of tools and
a more comprehensive, integrated approach than the Agency has typically utilized in
the past.

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APPENDIX A: SPECIAL TERMS AND THEIR USE IN

THIS REPORT

Ecosystem: A dynamic complex of plant, animal, and microorganism communities and the
non-living environment, interacting as a system.

Ecosystem functions or processes: The characteristic physical, chemical, and biological
activities that influence the flows, storage, and transformation of materials and energy within
and through ecosystems. These activities include processes that link organisms with their
physical environment (e.g., primary productivity and the cycling of nutrients and water) and
processes that link organisms with each other (e.g., pollination, predation, and parasitism).

Ecosystem Services: Those ecological characteristics, functions, or processes that directly
or indirectly contribute to the well-being of human populations or have the potential to do so
in the future.

Value: This term is used broadly to include contributions to human well-being and goals or
ends, such as social and civil norms (including rights) and moral, religious, and spiritual
beliefs and commitments.

Valuation: The process of measuring the value or the change in value in terms of the
contribution to a specified goal (e.g., human well-being, biodiversity conservation).

Valuation Method: An approach based on theory and data, for measuring the value or
change in value in terms of the contribution to a specified goal.

Monetary Valuation: Valuation in which the measurement is done in dollars or some other
financial unit.

Willingness-to-Pay (WTP) or Willingness-to-Accept (WTA) Valuation Methods:

Methods that estimate the trade-offs individuals are willing to make, expressed in monetary
terms. These approaches typically focus on the amount of money an individual is willing to
forgo to enjoy a positive change in terms of availability, quantity, or quality of the good or
service (willingness to pay). Alternatively, willingness to accept is the amount of monetary
compensation a person would accept in lieu of receiving that change.

Social-Psychological Valuation Methods: Methods that focus on individuals' or groups'
judgments of the relative importance of, acceptance of, or preferences for changes in
ecosystems, their components, or the services they provide. These methods typically focus on
choices or ratings among alternatives. Individuals making the judgments may respond on
their own behalf or on behalf of others (society at-large or specified sub-groups) and the
basis for judgments may be changes in individual welfare, changes in group welfare, or civic,
ethical, or moral obligations relevant to ecosystems and ecosystem services.

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APPENDIX B: DISCUSSION OF METHODS

BIOPHYSICAL RANKING METHODS

Method

I'olill ol oiilput Minis''

W lial is method unaided in

Source of Information About Value





measure'1

Does nielliod measure obser\ed heha\ ior.
\ erbal or u rilien expressions. or progress
relaled lo pre\ ioiisK identified uoal'.'

Wlio

expresses
\ nine"

Conservation
Value Method

Map of biodiversity, scarcity
and/or conservation values across
landscape

Contribution to biodiversity

Measurements related to previously
identified goal of biodiversity

Expert -
ecologist or
conservation
biologist



Deuree lo w Inch Method
1 las 1 Seen 1 )e\ eloped or
I lili/ed

Recommendations lor Research lo
Sireiiuihen I se of Method

I'oleniial for I'niiire I se b\ I P X in an Integrated
and 1 Apaiided \pproach for Valuation

Issues lu\ ol\ ed in
Implementation

Conservation
Value Method

Components of approach
used by

•	U.S. Department of
Agriculture,

•	U.S. Forest Service,

•	U.S. Fish and Wildlife
Service,

•	National Park Service,

•	Bureau of Land
Management,

•	IUCN,

•	The Nature
Conservancy,

•	Nature Serve

•	Integration of stakeholder ehcitatioii
approaches (e.g. social scientific
surveys) with ecological condition
mapping

•	R&D to show how GIS-based
systems could be designed to
integrate monetized and other
quantitative valuation approaches on
a common spatial and temporal GIS
background

•	Where sufficient data does not yet
exist, additional resources will need
to develop this information in order
to complete the methodology

•	L se to focus a\ ailable consen alion funds
related to conservation goals

•	Use as a prediction of ecological impacts that
would then be used as an input in an
economic valuation study

•	Use in combination with other non-monetary
value information (for example, from social-
psychological surveys) to characterize
preference-based values when monetization is
not possible or desirable

•	Use as a means of quantifying biophysical
impacts when they cannot be quantified (as
required by the OMB Circular A-4)

•	Issues w ilh the lack
of data

•	Currency and
confidence in
available data

•	Access to 'sensitive'
data represent
potential obstacles fo
the application of this
method

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Conservation Value Method

Overview In many contexts, decision makers need to know the conservation values for specific biophysical characteristics
across different geographies, and the distribution of these values across the landscape. Examples requiring the use of these values
include the need to know what sites are important for the conservation of biological diversity, and numerous decisions regarding the
protection of wetlands and mitigation of wetland impacts. Every landscape can be characterized by a suite of ecological properties
that form the basis for environmental, social, and economic values. The Conservation Value Method is a scientific process to map
these values across the landscape for use in decision making. Conservation value can be defined as a measure of the contribution of a
landscape unit to the conservation of species diversity, as defined or estimated by relevant experts.

This method also allows the incorporation of social preferences through the development of preferred conservation goals for
different biophysical and ecological properties. More than one set of goals can be developed to represent the interests and objectives
of different stakeholders. The conservation values are used as the basis for the evaluation of alternative actions in contributing to the
social goals that are being addressed. If the social goal is biodiversity conservation, for example, the evaluation of any action is a
measure of the contribution of this action to sustained ecosystem diversity and integrity.

This method assigns a value to each individual land area within a given region based on its contribution to a conservation-
based goal. This application of scientific information and methodology results in the mapping and valuation of biological and
ecological features in a regional context. This provides spatial value attributes for the representative biological and ecological
characteristics and features of that area. These can include both biotic factors (e.g., distribution and abundance of plant and animal
species) and abiotic factors (e.g., soils, hydrology, climate) that are spatially distributed across the landscape. Some of these features
in turn provide information about the ecosystem services provided by the land. This method can be completed with current
Geographic Information System-based technologies.

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Because each land area has multiple ecological dimensions, the values associated with the contributions of these different
dimensions are often weighted and aggregated, with the weights determined by the relevant stakeholders in a given decision context.
Different stakeholders will apply different weights, depending on the objective of their analysis (e.g., biodiversity vs. wetlands
protection). In addition, spatial information about ecological characteristics can be overlain with other spatial data of interest to these
stakeholders.

This process of weighting and mapping the resources that represent what people want to preserve is sometimes referred to as
"green printing." For example, groups such as Trust for Public Lands use this phrase when working with Watershed Stakeholder
groups to get them focused on steps to implement conservation. It allows for an effective approach with multiple stakeholders to
prioritize parcels in the landscape for acquisition and conservation.

Brief description of the method The Conservation Value Method, as detailed by Grossman and Comer (1994), was developed
as a general approach to create biodiversity-based conservation values. It represents a structured set of steps for constructing those
values, and is built to incorporate the input of stakeholders at multiple points in the process. These values are generated from system
attributes for uniqueness, irreplaceability, level of impediment, and ecological services.

The method begins with an identification of the species, ecosystems, and associated ecological services - and an assessment of
their status and condition across the landscape of concern. The evaluation is based on characteristics such as rarity, representation,
threat, landscape integrity, and other relevant factors. There are several national databases that can provide much of the baseline
information. The network of state Heritage Programs develop and maintain status and distribution information about thousands of
plants and animals, along with different vegetation and ecosystem types. The Integrated Taxonomic Information System (ITIS)
maintains a standardized list of species names for use by scientists and federal agencies. The U.S. Fish and Wildlife Service maintains
information about endangered species and wetlands, the U.S. Geological Survey manages databases characterizing ecosystem
characteristics and integrity, and the Department of Transportation manages information on the density and location of roads and

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infrastructure across the country. The standardized integration of these datasets within the Conservation Value methodology provides
a robust foundation for decision making.

The places where a given element of conservation interest is found (termed an "occurrence") is assigned a quality and viability
score based on attributes of size, condition, and landscape integrity. The trends and condition for each conservation element are
presented in a summary status attribute, a conservation rank (reference NatureServe, IUCN). The global assessment and the quality
information about individual occurrences are then used to develop a spatial "ecological value layer," which portrays a spatial
distribution of the conservation value along with metadata regarding the quality and confidence of each occurrence. This layer can
reflect the specific conservation goals of the stakeholders, as they can alter the relative importance of different conservation elements
based on their management or conservation objectives. To the extent that stakeholders are interested in multiple ecological features
(e.g., multiple species), the information for each ecological value layer is aggregated to create an overall "conservation value
summary." This summary value layer provides a spatially aggregated representation of the biodiversity and conservation values that
represent the values of the conservation or management stakeholders. The final (aggregate) conservation values are used to support
decision making, e.g., to prioritize preservation-based land acquisitions, mitigate wetland loss, direct point and non-point source
permits, etc. These spatial conservation values can also be integrated with socio-economic and other spatial data to integrate those
data into the decision-making process.

The Conservation Value Method was developed primarily to identify priority areas and activities that would sustain or improve
the condition of biodiversity and ecosystem health. This GIS based methodology can support different types of decisions by adding
different data and values to the model. For example, one could quantify Bureau of Land Management land for its value as recreational
use, natural resource extraction (timber, mineral, oil and gas), and water quality (denitrification, water purification) and quantity (flood
control, snow pack).

This method is often used to evaluate the impact of a proposed action on current conditions. This requires the development of
future scenario maps that can reflect a new policy, a development action, modeled population growth, a natural disaster, or any

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number of different change scenarios. The intersection of the change scenario with the conservation value model allows for clear
reporting on the changes to either the composite conservation value or the individual conservation values. This is often used to choose
between change scenarios (e.g., road placement, point source licenses), and to protect against potential threat (toxic transport, oil line
placement).

The Conservation Value Method can contribute to EPA decision making in a number of ways. First, in contexts where the
Agency's goals are defined in terms of conservation objectives or requirements, such as under the Endangered Species Act, the
method could provide a means of making decisions about where to focus available conservation funds. In addition to contributing to
decision making focused on specific conservation goals, the outputs from the conservation method could play a key role in EPA
decision making (and the C-VPESS valuation framework) in the following ways: a) it could be used as a prediction of ecological
impacts that would then be used as an input in an economic valuation study; b) it could be combined with other non-monetary value
information (for example, from social-psychological surveys) to characterize preference-based values when monetization is not
possible or desirable; and c) it could be used as a means of quantifying biophysical impacts when they cannot be quantified (as
required by the OMB Circular A-4).

Status as a method The Conservation Value Method approach represents a sequence of iterative steps that have been
developed by the scientific community over the past thirty years. (References?)The components that have been aggregated into this
emerging methodology include ecological classification and mapping standards, conservation ranking standards, conservation
planning methodology, and occurrence mapping standards. There is widespread use of various components of these methods across
U.S. federal agencies, though the utility use of the comprehensive integrated methodology has only recently become accessible and
manageable for the non-specialist. The ranking methodologies for conservation elements (plant, animals, and ecosystems) has been
documented in the scientific literature over many years and is in common use by numerous federal agencies (e.g., U.S. Department of
Agriculture, U.S. Forest Service, U.S. Fish and Wildlife Service, National Park Service, and Bureau of Land Management).
(References?) The viability and quality ranking criteria for the occurrences of conservation elements has been the topic of widespread

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analysis by IUCN, The Nature Conservancy, NatureServe and others. The conservation planning methods have emerged from
Australian natural resource agencies (e.g., CSIRO) and are well published in the conservation science literature. (References?) EPA
has used different components of this methodology to identify and prioritize rare and threatened species that need protection (e.g.,
working with the pesticide industry to protect biological diversity) and to characterize different wetland ecosystems to prioritize
protection activities. (References?)

This methodology is increasingly being used by the larger planning community for different purposes at multiple scales. The
examples listed below will illustrate the breadth of these applications. The Land Trust of Napa County has used the methodology to
identify priority conservation acquisitions for the next ten years. The U.S. Forest Service is testing its use for the development and
monitoring of National Forest plans. The Conservation Trust of Puerto Rico has applied these methods to clarify conservation and
development priorities and options across the island. The state of Mata Grosso in Brazil is using this approach to integrate a
conservation reserve program into private landholdings.

Decision contexts where this method could be used by EPA include:

•	Enumeration of biodiversity protection implications that result from policy changes (i.e., change of protection status for
isolated wetlands)

•	Identification of critical riparian habitat

•	Prioritization of remediation action on superfund sites

•	Due diligence reviews and Environmental Impact Statements as a prerequisite for permitting

•	Identification of reference conditions for establishment of baseline quality metrics for wetland and aquatic habitats

•	Assessment of the status of target species and ecosystems

•	Analysis of mitigation equivalencies and priorities

•	Baseline information for ecosystem integrity and environmental impact monitoring

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Strengths/Limitations

Conceptual Strengths/Limitations The Conservation Value Method will create a quantitative spatial representation of
ecological and biological values within a regional context. The spatial range of these analyses can vary from local to regional scales.
This data provides a baseline for a broad range of natural resource assessment and management decisions, and can be integrated with
spatial monetary valuations to inform cost-effective land management and regulatory decisions. The specific decisions will determine
that types of data and analyses that are required to address the question.

The Method's Strengths

•	The method is adaptable to address different questions.

•	The method can be run repeatedly to represent temporal change or different landscape scenarios.

•	Results are commonly aggregated to derive a single benefits number, but all of the native data is constantly maintained
in the system and can be presented separately.

•	The output is both understandable and communicable to the interested audience and other stakeholders. Provides the
opportunity for visualization of outcomes that many other methods lack.

•	The results are repeatable, and the process and algorithms are very transparent.

The method's weaknesses Issues with the lack of data, the currency and confidence in available data, along with access to 'sensitive'
data represent potential obstacles for the application of this method. There are many ways to create surrogate datasets that will allow
users to adapt to different types of barriers. Some training and tools are also required to use this method.

Practical Strengths/Limitations

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The assumption is that there is sufficient coverage of standardized biodiversity data required to implement these methods. The
standards for each step of the method have been developed, and the data that is required will be dependent upon the specific
application questions. Where sufficient data does not yet exist, additional resources will need to develop this information in order to
complete the methodology. In some cases, surrogate information and models are required to incorporate the spatial representation of
poorly inventoried conservation targets across the landscape.

This method requires local scientific data, knowledgeable scientific interpretation and conservation planning expertise. The
magnitude of the need is contingent upon the application and the current state of data and knowledge. There are many sources
available from which to obtain this knowledge.

Treatment of Uncertainty There are confidence measures built into the methodology that can be integrated into the decision-making
analysis or displayed independently for consideration. The most significant sources of uncertainty in the use of this method include:

•	The variability in the quantity and quality of the data

•	The limitations of scientific understanding of distribution and quality criteria for some ecological factors

•	The level of stakeholder understanding of the linkages between ecological components and the services they value

Research needs There is both a need and an opportunity to actively explore integration of stakeholder elicitation approaches (e.g.,
social scientific surveys) with ecological condition mapping. Additional R&D to show how GIS-based systems could be designed to
integrate monetized and other quantitative valuation approaches on a common spatial and temporal GIS background could yield
significant benefits.

Key References

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Brown, N., L. Master, D. Faber-Langendoen, P. Comer, K. Maybury, M. Robles, J. Nichols, and T. B. Wigley. 2004. Managing
Elements of Biodiversity in Sustainable Forestry Programs: Status and Utility of Nature Serve's Information Resources to
Forest Managers. National Council for Air and Stream Improvement Technical Bulletin Number 0885.

Grossman, D.H. and P.J. Comer. 2004. Setting Priorities for Biodiversity Conservation in Puerto Rico. NatureServe Technical Report.
Riordan, R. and K. Barker. 2003. Cultivating biodiversity in Napa. Geospatial Solutions.

Stoms, D. M., P. J. Comer, P. J. Crist and D. H. Grossman. 2005. Choosing surrogates for biodiversity conservation in complex
planning environments. Journal of Conservation Planning 1.

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Rankings Based on Energy and Material Flows
Introduction

Energy and material flow analysis is the quantification of the flows of energy and materials through complex ecological or economic
systems, or both. These analyses are based on an application of the first (conservation of mass and energy) and second (entropy) laws of
thermodynamics to ecological-economic systems. A recent report by the National Research Council (NRC) covers the basic elements and need
for such analyses (Committee on Materials Flows Accounting of Natural Resources, Products and Residuals 2004). The NRC report concludes
that information about material flows can be a very useful input for policy decisions. It can be used to identify potential environmental
concerns and key sources of pollution and to develop strategies for preventing environmental releases.

This section provides general background on energy and material flow analysis as a means of identifying and quantifying important
relationships within ecological and economic systems. It then discusses two methods that translate the physical energy and material flows into
measures that could be used in the context of ecological valuation. The first is embodied energy analysis, which estimates the direct and
indirect energy (or more correctly, available energy or "exergy") cost of goods and services. The second is ecological footprint analysis, which
estimates the biologically productive land or water areas required (directly or indirectly) to meet various consumption patterns. We also briefly
discuss the use of the concept of "emergy" for estimating energy costs and valuation.

Energy and Material Flows Analysis

Energy from the sun drives plant productivity as well as climate and hydrologic cycles, nutrient cycles, ocean currents, weathering and
soil formation. Thus a study of energy and material flows in ecosystems relates very directly to the production of ecosystem services.
Ecologists have long utilized studies of the flow of energy and materials (e.g., nitrogen, phosphorus) through ecosystems as a way of describing
key relationships and understanding the functioning of those ecosystems. Early studies of energy flow in aquatic (e.g., Lindeman 1941) and
terrestrial (e.g., Golley 1960) systems illustrated how energy moved through food chains. Ground-breaking analyses of the cycling of critical

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nutrients in lakes (e.g., Hutchinson 1947) and forests (e.g., Likens and Borman 1977) set the stage for many subsequent analyses and
established the field of biogeochemistry. Studies of energy and materials flows can be especially useful for understanding how changes to an
ecosystem, such as an increased or decreased level of pollution, may alter the system and the services it provides. For instance, increases or
decreases in the inputs of nitrogen to a forest from acidic deposition may impact forest productivity, species composition, and nitrogen runoff in
streams and rivers (e.g., Johnson and Lindberg 1991). Larsson, et al. (1994) used energy and material flows to demonstrate the dependence of a
renewable resource such as commercial shrimp farming on the services generated by marine and agricultural ecosystems. The committee
seconds the view expressed by the NRC (2004) that analyses such as these can provide very valuable information about ecological services and
how the flow of services might change in response to specific stressors.

The energy and environmental events of the 1960s and 1970s prompted a number of economists, ecologists, and physicists to examine
the energy and material flows underlying the economic process (Boulding 1966, Georgescu-Roegen 1971, 1973). Ecologists noted the
importance of energy in the structure and evolutionary dynamics of ecological and economic systems (Lotka 1922, Odum and Pinkerton 1955,
Odum 1971). The integration of the first law of thermodynamics with the economic system was first made explicit in the context of an
economic general equilibrium model by Ayres and Kneese (1969) and subsequently by Maler (1974). It is also a feature of a series of linear
models developed after 1966 (Cumberland 1966, Victor 1972, Lipnowski 1976). All reflect the recognition that the earth is a
thermodynamically closed (but not isolated) system, with energy from the sun crossing the boundaries and maintaining the structure and
function of the earth system. A closed system must satisfy the conservation of mass condition. Ayres (1978) described some of the important
implications of the laws of thermodynamics for the economic production process, noting that both manufactured and human capital require
materials and energy for their own production and maintenance (Costanza 1980).

A key feature of energy flow analysis is the recognition of the importance of energy quality, namely, that a kcal of one energy form
(e.g., electricity) may produce more useful work than a kcal of another (e.g., oil). Estimating total energy consumption for an economy is
therefore not a straightforward matter because not all fuels are of the same quality, that is, they vary in their available energy, degree of
organization, or ability to do work. This effort to incorporate energy quality is often referred to as "second law analysis."

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Embodied Energy Analysis

As noted, methods have been developed that seek to use energy and material flows information to determine values associated with
different systems or changes in those systems. One such method is embodied energy analysis. The embodied energy method assesses the direct
and indirect energy costs of economic and ecological goods and services. It uses input-output tables to determine the direct and indirect energy
inputs used to produce these goods and services. Although there is no stated Agency policy to use or develop supplemental valuation
methodologies in this area, there is substantial Agency interest in how Energy and Material Flow methods might aid decision making. Recent
efforts to explore the utility of such methods, mostly at the regional or local level, are underway (Bastianoni et al., 2005, Campbell 2001, 2004,
Lu, et al. 2006).

Some ecologists and physical scientists have used estimates of embodied energy to implement an energy theory of value either to
complement or replace the standard neoclassical theory of subjective utility-based value (Soddy 1922, Odum 1971, 1983, Slesser 1973,
Gilliland 1975, Costanza 1980, Cleveland, et al. 1984, Hall, et al. 1992). The energy theory of value is based on thermodynamic principles,
where solar energy is recognized to be the only primary or external input to the thermodynamically closed global ecosystem. At the global
scale, the traditional primary factors of production (labor, manufactured capital, and natural capital) are viewed as intermediate factors
(Costanza 1980).

There has been ongoing debate about the validity of an energy theory of value (Brown and Herendeen 1996). Some believe that it is the
only reasonably successful attempt to operationalize a general biophysical theory of value that does not hinge completely on consumer
preferences (see also Patterson 2002). Neoclassical economists, on the other hand, have criticized the energy theory of value as an attempt to
define a concept of value that does not directly reflect consumer preferences regarding the good being valued (see Heuttner 1976). This
criticism is, on the one hand, axiomatic, since a major purpose of an energy theory of value is to establish a theory of value not completely
determined by individual preferences. On the other hand, techniques for calculating embodied energy utilize economic input-output tables.
These tables summarize production interdependences, but they are not completely independent of consumer preferences, which helped to
structure the production interdependences over time. Neoclassical economists also question the primary status of energy, because in any

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concrete, short-term situation the scarcity and prices of the conventionally-defined inputs of manufactured capital, labor, and technology are
also important. While not denying the importance of these short-term considerations, energy theorists take a broader, more evolutionary
perspective, recognizing that these factors are intermediate and that production relationships adapt over time.

As noted, the energy theory of value (like the labor theory of value developed by classical economists) is inherently based on relative
production costs, i.e., it yields a measure of (direct plus indirect) energy cost. The question arises as to when these energy-based production
cost estimates can provide a measure of value. This is similar (but not identical) to the question that arises in the context of replacement costs
based on the standard economic concept of opportunity cost (see Chapter 4.1.7 and more detailed discussion of replacement costs in the
Appendix below). In economic systems, marginal cost and price will be equal in a perfectly competitive equilibrium. This means that, in the
absence of other market distortions, an estimate of marginal cost can provide a proxy for the value of an additional unit of production.

Similarly, an estimate of production cost can provide a proxy for value, but only under certain circumstances (see discussion in section on
replacement cost). For example, the aggregate individuals must be willing to incur these costs rather than forego the good or service. One
difference between replacement and production costs is that while replacement costs are hypothetical, production costs have already been
incurred, implying that aggregate individuals were willing to incur the costs, thus satisfying this condition. To the extent that the necessary
conditions are met, energy costs can provide information about the value of the associated goods or services as defined by the energy theory of
value.

Costanza, et al. (1989) provide an example of wetlands valuation that uses both a conventional WTP approach and a simplified energy
analysis approach based on the gross primary productivity (GPP) of coastal wetlands in Louisiana. The energy analysis valuation technique
compared total biological productivity of a wetland versus an adjacent open water ecosystem. Primary plant production, which supports the
production of economically valuable products such as fish and wildlife, was converted to a monetary value based on the cost to society to
replace this energy source with fossil fuel as measured by the overall energy efficiency of economic production. While the results of the WTP-
and GPP-based methods were fairly consistent, the authors note that the GPP approach probably represented an upper bound and "may

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1	overestimate their value if some of the wetland products and services are not useful (directly or indirectly) to society" (Costanza, et al. 1989, p

2	341). However, it should be noted that the basic assumptions underlying an energy theory of value imply that there is no reason to expect

3	measures based on energy cost to be the same as preference or WTP-based measures of value.

4	Ecological Footprint Analysis

5	The ecological footprint (EF) method is a variation of energy and material flow analysis that converts the impacts to units of land or

6	water rather than energy or dollars. The EF for a particular population is defined as the total "area of productive land and water ecosystems

7	required to produce the resources that the population consumes and assimilate the wastes that the population produces, wherever on Earth that

8	land and water may be located" (Rees 2000). While usually discussed in the context of the footprint of specific human populations, this

9	concept can also be applied to non-human populations. For example, a portion of the southern Chesapeake Bay has been set aside as a blue

10	crab sanctuary since large numbers of the organisms spawn in this area relative to elsewhere (Virginia Marine Resources Commission, Newport

11	News, VA). In the context of human societies, input-output methods (see previous discussion) are used to estimate direct and indirect land

12	requirements.

13	Although there are ongoing debates about specific methods for calculating the ecological footprint (Costanza 2000, Herendeen 2000,

14	Simmons, et al. 2000), the ecological footprint is an effective device for presenting current total human resource use in a way that

15	communicates easily to a broad range of people (http ://www.footprintnetwork. org/). In terms of valuing ecosystem services, the ecological

16	footprint concept is most useful as an index of the quantity of ecosystem services consumed (expressed in units of a standardized land area) for

17	various consumption patterns. This measurement, however, does not directly convert to a monetary measure of the value of ecological services.

18	It does, however, allow a relative comparison of one footprint to another based on areas or sizes involved. Under this approach, ceteris paribus,

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a population that has a smaller footprint is viewed as more sustainable. On the other hand, a larger footprint implies a larger biocapacity
supporting a given population and a larger required contribution of ecosystem services to maintain that population in its current state.

Emergy Analysis

Emergy analysis shares many of the same goals and assumptions as embodied energy analysis. For example, solar emergy is defined as
"the available solar energy used up directly and indirectly to make a service or product" (Odum 1996). Emergy analysis differs from embodied
energy analysis and ecological footprint analysis in terms of the method used to estimate the energy required. While embodied energy and
footprint analysis use methods based on input-output (a well-developed set of methods for this type of accounting), emergy analysis uses
different methods (See recent work by Ukidwe and Bakshi, in press).

Emergy analysis starts with the creation of an energy flow diagram. The "Solar Transformity" is then defined as "the solar emergy
required to make one Joule of a service or product" (Odum 1996). This is calculated by dividing any flow in the diagram by the total solar
energy input. Odum and coworkers have thus calculated the emergy of the earth's main processes, such as the total surface wind, rain water in
streams, the sedimentary cycle, and waves absorbed on shore, to be that of the total emergy input to the earth (Odum 1996). Each of these
processes is assigned the total value of incoming sunlight because they are considered co-products of the global geological cycle and cannot be
produced independently with less amount of the total emergy.

However, emergy has encountered considerable resistance and criticism, particularly from economists, physicists, and engineers (Hau
and Baksi 2004, Ayres 1998, Cleveland, et al. 2000, Mansson and McGlade 1993, Spreng 1988). Consequently, the emergy approach has only
been used by a small circle of researchers, although some work at EPA is ongoing (U.S. Environmental Protection Agency 2005). Emergy's
accounting method does not produce an estimate of the energy cost of goods and services, but rather "the relative equivalence between energies
of different kinds in terms of a universal quality factor." This concept is difficult to understand and to apply in a standard accounting
framework. Although the committee as a whole did not study the debate over emergy in detail, the committee believes that substantial
questions exist regarding the appropriateness and usefulness of emergy as a method for valuing ecological systems and services.

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Key References

Golley, F.B. 1960. Energy dynamics of a food chain of an old-field community. Ecological Monographs 30: 187-206.

Hutchinson, G.E. 1947. A direct demonstration of the phosphorus cycle in a small lake. Proceeding of the National Academy of Science 33:
148-153.

Johnson, D. W., and S. E. Lindberg (eds.). 1991. Atmospheric Deposition and Forest Nutrient Cycling: A Synthesis of the Integrated Forest

Study. Ecological Series 91, Springer-Verlag, New York. 707 p.

Likens, G.E and F.H. Borman. 1977. The Biogeochemistry of a Forested Ecosystem. Springer-Verlag, New York. 154 p.

Lindeman, R.E. 1942. The trophic-dynamic aspect of ecology. Ecology 23: 399-418.

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ECOSYSTEM BENEFIT INDICATORS

Method

I'ol'lll ol oiilput Minis''

W lial is melliod intended in

Source of Information Abonl Value





measure'

Does melliod measure ohser\ed heha\ ior.
\ erbal or u rilien expressions, or progress
relaled lo pre\ ioiisIs idenlil'ied uoal'.'

Who

expresses
\ alne'.'

Ecosystem
Benefit Indicators

Map of the supply of
ecosystems/services showing
quantities of expressed or
estimated demand for those
ecosystems/services across a
landscape

Quantitative but not
monetary approach to
preference weighting for the
ecological effects of policy
options

Measurements related to demand variables
that can be identified by experts or non-
expert lay publics and supply variables as
identified by experts.

Expert and
selected
non-expert
lay public

2



Decree lo u hicli Melliod 1 las

Recommendations lor Research lo

I'oleniial for I'nlnre I se b\ I P-\ in an

Issues ln\ol\ed in



I'.een l)e\eloped or I lili/.ed

Sireiiizihen I se of.Method

Inieuraled and Lxpanded \pproach

Implementation







for Valuation



Ecosystem

The method is new and

• Integration of EBIs with biophysical

• Input to a wide variety of trade-

• Do not directly yield

Benefit

relatively undeveloped

endpoints

off analyses (for regulatory

dollar-based

Indicators



• Integration of EBIs with econometric

analyses or performance

ecological benefit





valuation methods (benefit function

measures)

estimates





transfer, stated preference and choice

• Use as part of public processes

• Do not in themselves





modeling)

designed to communicate the

weight or estimate





• Suitability for group decision

implications of a change or policy

the trade-offs





techniques, such as mediated modeling

across a variety of scales

associated with





• Practical application to illustrate data

• Use as inputs to economic and

different factors





needs and measurement issues

econometric methods such as

relating to benefits







benefit transfer, or stated

• Uncertainty with







preference models

regard to how







• Use to systematize alternative

indicators are







choice scenarios in choice

perceived,







experiments and stated preference

particularly when







surveys

presented visually









should be









acknowledged

3

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Introduction

This report describes a range of valuation methods. The choice of method will depend on the environmental question at hand, the
political and regulatory process involved, and differing philosophical perspectives on the nature of value and how it is to be determined by
society. All of these methods, however, require the analyst or decision maker to be informed.

Two basic forms of information are required: the knowledge of what is at stake in nature and the ability to determine how ecological
endpoints change as a result of management or regulation. The first piece of information comes under the realm of biophysical production
function analysis. If the agency can achieve clear, actual or predicted production function-based outcomes, that would be a great advance over
current practice.

Assuming these kinds of information and analysis are available, social scientists are then called upon to weight, prioritize, or value
different outcomes in nature. What kind of information should be relied upon for weighting, priority-setting, and valuation of ecological
changes?

Recommendation: The committee advocates the Agency more broadly collect and communicate ecosystem benefit indicators (EBIs) to
inform the social weighting and valuation of ecosystem services. EBIs are not themselves a valuation method. Rather they are an inventory of
data and set of principles that should be used to inform the public or analyst as part of any valuation exercise.

Elsewhere in this report the committee has emphasized the importance of ecosystem services' spatial and landscape context. Where
services arise is very important, both ecologically and socially. From a social science standpoint, the determinants of value depend upon the
landscape context in which ecosystem services arise. Habitat support for recreational and commercial species, water purification, flood damage
reduction, crop pollination, and aesthetic enjoyment are all enjoyed in a larger area surrounding the ecosystem in question. EBIs allow for
spatial representations (both geo-coded data and corresponding visual depictions) of social and biophysical features that enhance or decrease
the benefits of a particular ecosystem services in particular places.

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Regulatory and ecological ecosystem assessments, including many of those reviewed by this committee, often ignore information that is
fundamental to valuation - however valuation is defined. For example, how many people benefit from a particular ecological function or
service? The number of people who can enjoy the service in a given location is an example of an important EBI.

•	The committee also found scant evidence that the Agency analyzes the scarcity of particular ecosystem services, the presence of
substitutes for those services, or the dependence of environmental benefits on the presence of complementary goods and
services. EBIs are a way to relatively quickly and cheaply address this information gap.

•	EBIs are of practical use to the Agency because the cost of collecting them is relatively low. EBIs are generated from GIS data
and can be quickly assembled, usually using existing data sets employed by federal, state, and local governments.

•	EBIs can and should be used to educate decision makers and stakeholders about the underlying complexity of ecological and
economic relationships. They are not a way to simplify the decision maker's problem. Rather, they provide basic information
that informs the decision process about the trade-offs arising from a particular decision.

Examples

To illustrate the use and benefits of EBIs, consider the following example: wetlands can improve overall water quality by removing
pollutants from ground and surface water. This ecological function is valuable, but just how valuable? To answer this question one can count
a variety of things, such as the number of people who drink from wells attached to the same aquifer as the wetland. The more people who drink
the water protected by the wetland, the greater its value.

But other things matter as well. For example, is the wetland the only one providing this service or are others contributing to the
aquifer's quality? The more scarce the wetland, the more valuable it will tend to be. There may also be substitutes for wetland water-quality
services provided by other land-cover types such as forests. Mapping and counting the presence of these other features can further refine an
understanding of the benefits being provided by a particular wetland.

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Many ecosystem benefits arise only in the presence of complementary features. Recreation typically requires access to natural areas.
Road, trail, dock or other forms of access are thus important to the analysis of benefits. In some cases, if there is no access, there can be no
benefit.

Consider another type of environmental benefit: aesthetic value arising from natural viewscapes. Here, relevant stakeholders and
decision makers would benefit from the following kinds of EBI: population in viewshed of the natural area (primary demand); percent of that
population's viewshed that is natural (scarcity); the number and extent of substitute viewsheds for this population (substitutes); the presence of
roads, trails, boatable surface waters, public lands, and access points that allow the natural area to be viewed (complements).

In general, EBIs should be specific to the ecosystem service and benefit in question. Consider two different ecosystem benefits:
recreational angling and provision of clean drinking water. The EBIs relevant to these two benefits will be different. In both cases, the number
of people benefiting is relevant, but the populations are different. Demand for recreational angling would involve assessment of the number of
potential anglers. This population is different from the population benefiting from a given aquifer's water quality. The determination of
scarcity and substitutes is very different as well. All of these examples of EBIs can be mapped and counted using geo-coded social (e.g.,
census) and biophysical data.

Brief Description

EBIs are countable landscape features that tell us about demand for, scarcity of, and complements to particular ecosystem services.
Ecosystem benefit indicators (EBIs) are quantitative inputs to valuation methods. They can serve as important inputs to valuation methods as
diverse as citizen juries and econometric benefit transfer analysis, which is a monetary weighting technique. EBIs provide a way to illustrate
ecological benefits in a specific setting. For example, if water is available at a particular place and time, how many water users (e.g.,
recreationists, farms) are present to enjoy that service? What other sources of water are available to those same users? These questions are
central to economic valuation of the resource.

Key inputs - EBIs are drawn mainly from geospatial data, including satellite imagery. Data can come from state, county, and regional
growth, land-use, or transportation plans; federal and state environmental agencies; private conservancies and nonprofits; and the U.S. Census.

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Key outputs - Spatially specific measures (both geo-coded data and corresponding visual depictions) of social and biophysical features
that enhance or decrease the desirability of particular ecosystem services.

Scale - The method is entirely scalable. One strength, however, is the ability to relate ecological and economic features in a specific
landscape context. For example, the method can be applied to individual projects, investments, or decisions made in a particular watershed.
They can also be expressed as local, regional, state, or national aggregates.

Example of How the Method Could be Used as Part of the C-VPESS Framework

The method relates to framework item (4): "Characterization of the Value of Changes in Monetary and Non-Monetary Terms." Benefit
indicators are countable features of the physical and social landscape. More specifically, they are features that influence - positively or
negatively - ecosystem services' contributions to human well-being. The consumption of services often occurs over a wide scale. For
example, habitat support for recreational and commercial species, water purification, flood damage reduction, crop pollination, and aesthetic
enjoyment are all services typically enjoyed in a larger area surrounding the ecosystem in question. EBIs help people understand the larger
social and physical landscape so that they can better assess the relative importance of particular services in particular places at particular times.

The value of ecosystem services is likely to be affected by the following factors: the ecosystem feature's scarcity, natural and built
substitutes, complementary inputs, and the number of people in proximity to it. For a given ecosystem service scarcity, substitutes,
complements, and demand can be related to landscape characteristics. Landscape features that relate to human well-being can be systematically
counted and mapped, and then aggregated into bundles of indicators (an index). Some indicators are biophysical, others relate to the socio-
economic environment.

Benefit indicators are an input to a wide variety of trade-off analysis approaches, but do not independently make or calculate the results
of such trade-offs. First, they can be used as ends in themselves as regulatory or planning performance measures. Second, they can be used as
part of public processes designed to communicate the implications of a change or policy across a variety of scales. Indicators or an index based
on them can then be used to elicit public preferences over environmental and economic options - as in mediated modeling exercises or more
informal political derivations. In this way, benefit indicators are a potentially powerful complement to group decision processes. Third, they

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can be used as inputs to economic and econometric methods such as benefit transfer, or stated preference models. This is an area where
research is needed. Economic methods must be developed to link indicator outcomes to dollar-based valuation in a way that is both statistically
and theoretically sound. In principle, benefit indicators could be used to calibrate the transfer function in benefit transfers. They could also be
used to systematize alternative choice scenarios in choice experiments and stated preference surveys.

As a method to inform the weighting of ecosystem services in a social decision context, the benefit indicators method requires
information provided by the biophysical sciences. The method requires spatially depicted biophysical endpoints. EBIs are then related to those
endpoints.

The method can be applied to any ecosystem service benefit where benefits are related to the spatial delivery of services and social
landscape in which the benefit is enjoyed. Existence benefits (where spatial location is irrelevant to both provision and value) are the only
ecosystem benefit category where the method would be inapplicable.

The data used in EBI analysis is well-suited to delivery via a national data bank.

Status as a Method

The method is new and thus relatively undeveloped. EPA has funded a small amount of research on the topic. For citations to peer
reviewed research, see below.

Strengths/Limitations

EBIs are designed to be a relatively non-technical way to express the factors that contribute to conventional economic measures of
benefits provided by ecosystem services. Their simplicity, and transparency, is an advantage. They can be used to communicate and educate.
By stopping short of monetary estimation of benefits (unless integrated in a benefit function transfer method) they are also a way for the agency
to overcome resistance to economic assessments of the natural world - while still conveying outcomes in a way designed to be consistent with
economic principles and the dependence of human well-being on natural assets.

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The principle disadvantage is that they do not directly yield dollar-based ecological benefit estimates. They also do not in themselves
weight or estimate the trade-offs associated with different factors relating to benefits (though as noted previously they can be married to more
formal methods designed to do such weighting).

Because indicators can be cheaper to generate than econometric value estimates they better allow for landscape assessment of multiple
services at large scales.

Treatment of Uncertainty

A core rationale for the use of a benefit indicator approach is to explicitly convey the sources of complexity - and hence uncertainty -
characterizing biophysical systems and the benefits arising from them. The visual depiction of benefit indicators, for example, can mimic
sensitivity analysis by presenting a range of benefit scenarios in GIS form. However, the visual depiction of quantitative information introduces
uncertainties of its own. In particular, visual depictions can strongly influence perceptions. Uncertainty with regard to how indicators are
perceived, particularly when presented visually, should be acknowledged.

Research Needs

•	Integration of EBIs with biophysical endpoints

•	Integration of EBIs with econometric valuation methods (benefit function transfer, stated preference and choice modeling)

•	Suitability for group decision techniques, such as mediated modeling

•	Practical application to illustrate data needs and measurement issues

Satisfying these needs would be a significant undertaking in terms of expertise, financial resources, and coordination within the agency.

References

James Boyd, "What's Nature Worth? Using Indicators to Open the Black Box of Ecological Valuation," Resources, 2004.

James Boyd and Lisa Wainger, "Landscape Indicators of Ecosystem Service Benefits," 84 American Journal of Agricultural Economics, 2002.

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1	James Boyd, Dennis King, and Lisa Wainger "Compensation for Lost Ecosystem Services: The Need for Benefit-Based Transfer Ratios and

2	Restoration Criteria," 20 Stanford Environmental Law Journal, 2001.

3	Lisa Wainger, Dennis King, James Salzman, and James Boyd, "Wetland Value Indicators for Scoring Mitigation Trades," 20 Stanford

4	Environmental Law Journal, 2001.

5

6	.

7

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MEASURES OF ATTITUDES. PREFERENCES. AND INTENTIONS

\lelhod l-'orni iil'iiMipui uiuls'.' \Vh;ii is melhoil niiciiclcd In incisure'.'

Source of liifnrnialioii Aboui \ alue







Does method measure obsened lvha\ ior.
\ erhal or u mien expressions, or progress
relaled lo pre\ ioiisK ideiililied izoal'.'

\\'Ik> expresses
MlllIC?

Survey questions
measuring social-
psychological
value constructs

Quantitative indices of attitudes ,
preference rankings, or
behavioral intentions toward
depicted environments or
conditions

Public concerns, attitudes, values,
beliefs, and behavioral intentions

Verbal reports, choices, rankings, ratings

representative
sample from
public

Conjoint attitude
survey questions

Indices of expressed attitudes or
preferences for multi-attribute
alternatives and implied trade-off
weights for composite attributes

Public concerns, attitudes, values,
beliefs, and behavioral intentions
related to specific trade-offs among
attributes

Choices, rankings, ratings

representative
sample from
public

Individual
Narratives

Summaries of individual's value-
relevant narratives

Implied knowledge, beliefs, attitudes
and concerns

Verbal reports from individual
stakeholders and lay publics

select sample
from public

Mental Models

Systematic, structured models of
beliefs and assumptions
underlying value positions

Value-relevant knowledge, beliefs
and assumptions and their
interrelationships

Verbal narratives from individual
stakeholders, lay publics or experts

select sample
from public

Behavioral
Observation/Trace

Observations of current or prior
(trace) use associated with
ecosystems/services

Responses to policies, outcomes, and
consequences, in situ

Current or past behavior associated with
changes in ecosystems/services

representative
sample of
current or past
visitors/users

Interactive
Environmental
Simulation
Systems

Observations of behavior in
simulated/game environments,
implied preferences

Responses to investigator-controlled
changes in simulated (virtual)
environments

Interactive behavior in response to changes
in simulated environments

representative
sample from
public

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Deurce id \x Inch \lollioil
1 las 1 '.ecu 1 )c\ eloped or
I lili/ed

Recommendations lor Research id
Sireimiheii I seoI'Melhod

Roleuiial lor I'ulure I se b\ IP \ in an
Inieuraled and 1 Apauded Approach lor
Valuation

Issues lu\ ol\ ed iu
Implementation

Surveys

Including

Attitude,

Preference and

Behavioral

Intention

Questions

•	Survey questions
measuring social-
psychological
constructs are the
oldest and most
frequently used
methods for
determining public
beliefs, concerns, and
preferences.

•	Survey questions have
been used, and
continue to be used,
effectively by all levels
of government to
measure citizen desires
concerns and
preferences.

•	How can allilude/preference surveys
best be used in EPA policy and
decision making, including how
decision makers can and should use
the relative quantitative (non-
monetary) value indices provided?

•	How can attitude/preference indices
be used to cross-validate decisions
implied by estimates of monetary
values (e.g., from CBA) for
alternative policies/outcomes?

•	How should value indices from
attitude/preference surveys be
integrated with bio-ecological and
economic value indices to
strengthen support for
ecosystems/services protection
policies and decisions?

•	Can contribute to initial problem
formulation by identifying ecological
services and impacts that most
concern citizens and/or identified
stakeholders, as well as by
uncovering assumptions, beliefs, and
values that underlie that concern

•	Can help to determine and quantify
socially important assessment
endpoints

•	Can be used to assess relative public
preferences among policy options
and their attributes

•	Quantitative attitude/preference
indices may be especially useful
when the values at issue are difficult
to express or to conceive in monetary
terms or where monetary valuations
are viewed as ethically inappropriate

•	Can be used to help inform and to
systematically involve publics in the
balancing of multiple values for
ecosystems/services protection
decisions

•	Institutional barrier of the
Paperwork Reduction Act

•	Responding public may not
have adequate knowledge
and understanding of
complex ecosystem
processes, or well-formed
opinions and preferences
for protection options

•	Designing and
implementing a well-
designed survey requires
expertise that may not be
sufficiently represented
within the Agency (see
Appendix C

Conjoint
Attitude
Survey
Questions

Relatively new variation on
survey methods used
sparingly in environmental
valuation context, but
increasingly being used in
business/marketing, health
care, tourism-recreation,
and other value assessment
applications

How do the overall values for multi-
attribute conjoint policy options and the
individual attribute weights inferred
from choices/ratings relate to separately
assessed values for the same policies
and attributes?

What are the specific advantages and
disadvantages of conjoint methods for
ecosystems/services value assessments?

• May be especially well-suited for
gauging public preferences across
sets of complex multi-dimensional
policy alternatives, as are likely to be
involved in many EPA regulations
and actions for ecosystems and
services protection

Same as for surveys in general,
plus special concerns about
publics' ability to understand
and respond appropriately to
complex, multi-attribute
alternatives

Special experimental designs
and data analyses can be
complex and are unfamiliar to
many researchers and analysts

Individual

• Infrequently applied in

• What productive roles can

• Can make important contributions to

The selection of participants can

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Degree to which Method
Has Been Developed or
Utilized

Recommendations for Research to
Strengthen Use of Method

Potential for Future Use by EPA in an
Integrated and Expanded Approach for
Valuation

Issues Involved in
Implementation

Narratives

EPA valuation
contexts, but
increasingly used to
address social,
psychological, and
anthropological
questions related to
values, attitudes, and
behavior intentions in
other environmental
management contexts

individual interviews and other
qualitative methods play in Agency
policy and decision making?

• How should the results of
qualitative analyses best be
integrated with quantitative
assessments (bio-ecological,
attitude/preference, and economic)
to strengthen support for
policy/decision making?

improving the design, development,
and pre-testing of more formal
surveys that can provide reliable and
valid quantitative assessments of
public concerns and values
• Can assist in identifying and
articulating the conceptual basis of
public values and concerns,
especially in early stages of problem
formulation and value assessment

have very important effects on
outcomes—formal
representative (probability)
sampling is not typically used
and no scientifically accepted
alternative selection method has
been developed.

Rigorous qualitative analysis
methods have been developed
but are rarely used and
qualitative methods in general
have not been adequately tested
in ecosystems/services valuation
contexts.

Mental Models

• A relatively new
variation on individual
narrative procedures
which can use rigorous
analytic methods to
extract and structure
participant's
knowledge, beliefs,
and assumptions into a
coherent logical
structure.

•	How might mental models be
effectively used to design
appropriate value elicitation and
assessment methods for ecosystems
and services?

•	How might mental model structures
best be integrated with the results of
other methods to provide deeper
insights into value assessment to
support policy/decision making?

•	Appropriate precursor (i.e.,
formative analysis) to any formal
survey or preference elicitation
method, to improve the validity and
reliability of the method.

•	May be especially useful for
exploring the bases of value conflicts
between segments of the public, or
between publics and expert/scientific
opinions.

Research and development is
needed to secure a consistent
and rigorous set of methods for
qualitative analysis and mental
model construction.

Behavioral

Observation/

Trace

• Relatively new and
untested in value
assessment and policy
formulation contexts,
but research and trial
applications are
increasing

• How might the development of
emerging behavior observation and
behavior trace methods be shaped to
effectively contribute to Agency
policy and decision making needs?

• Might be used to attain quantitative
measures of human use levels useful
in conjunction with economic
measures or as separate measures to
be correlated with changes in
ecological conditions

In-situ field observations are
constrained to existing
conditions and are subject to the
effects of uncontrolled
variables.

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Deuree id \x Inch Method
1 las 1 Seeii 1 )e\ eloped or
1 lili/ed

Recommendations lor Research id
Sireimtheii I se of Method

Roleutial for future I se b\ IP \ in an
Integrated and 1 Apauded Approach for
Valuation

Issues lu\ ol\ ed in
Implementation

Interactive
Environmental
Stimulation
Systems

• Relatively new and
untested in value
assessment and policy
formulation contexts,
but research and trial
applications are
increasing

• How might die development of
emerging interactive computer
simulation and game methods be
shaped to effectively contribute to
Agency policy and decision making
needs?

•	Can engage and commumcale Willi
public audiences about what
outcomes they prefer and policies
required to achieve those outcomes

•	Respondents can learn through
experience about how the
ecosystems/services of interest
respond to various policies or policy
aspects and can progressively modify
and test their expressed policy
preferences

•	Provides opportunities to introduce
and experimentally control policy
relevant (and confounding) variables
in evaluated policies

Technological demands can be
high, but off-the-shelf
environmental simulation and
VR systems are increasingly
available and affordable.
As simulations approach VR
standards, demand for detailed
specification of environmental
conditions and processes
increases, possibly exceeding
current bio-ecological
knowledge in some cases

1

2	EPA has a number of laws, regulations and guides to assure that "the Agency considers public concerns, values, and preferences when

3	making decisions" (EPA 2003, p. 1). The social-psychological methods described in this section are consistent with that goal and can also

4	contribute to systematic quantitative assessments of the values of protecting ecosystems and ecosystem services. Survey methods are the most

5	frequently used means for identifying public values and concerns ("what people care about") and for measuring the degree of public preference,

6	acceptance and support for alternative environmental outcomes and associated social consequences (see Appendix C for a detailed discussion of

7	survey methodology). Surveys are also used to predict how various segments of the public are likely to respond to projected changes in

8	environmental conditions and to alternative management means for affecting those changes. Additional methods, such as individual narrative

9	interviews, can support agency decision making by elaborating and enriching understanding of the different perspectives of various
10	stakeholders and concerned citizens.

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EPA's charge to protect ecosystems and ecosystems services is consistent with widely shared public concerns and values (e.g., Dunlap,
et al. 2000). However, the formulation and implementation of specific ecological protection policies will often involve scientific and technical
considerations that the lay public cannot be expected to fully understand and appreciate. Surveys and the other methods described in this
section have proven effective in uncovering assumptions, knowledge, beliefs, and feelings that underlie expressed preferences and concerns so
that decision makers can better understand and address conflicts between various publics, and between public preferences and ecological
science. Moreover, there are a number of methods for introducing relevant information into or prior to a systematic survey that can help to
assure that respondents have an adequate and appropriate foundation for expressing requested preferences and other judgments (see Appendix
C).

While public opinion is sometimes directly used to make policy decisions (see Chapter 4 and Appendix B sections on Referenda and
Initiatives and on Citizen Valuation Juries), social-psychological assessment methods more typically are intended for decision support. These
methods address the psychological foundations for subsequent actions toward the measured alternatives, including political support, direct,
indirect, or hypothetical monetary payments, and acceptance of and compliance with relevant regulatory mandates. Typically, separate
measures are reported for several different value dimensions (e.g., aesthetic, ethical, personal-utilitarian, civic) across designated sets of policy
alternatives or for specific features of those alternatives. Consistent with a multi-attribute value framework, there has been little emphasis on
mapping all expressed concerns and preferences onto a single, universal value scale (as required for economic cost-benefit analysis methods,
for example). Differences between different value dimensions or between various subsets of the public are not typically resolved through
aggregation algorithms or other calculation devices within the assessment process. Rather, resolution of such differences is more typically
deferred to later stages of the decision making process, where information integration, deliberation, and negotiation is left to authorized
decision makers or is addressed in more or less formal interactions between stakeholders/publics and decision makers.

The social-psychological approach to assessing the value of ecosystems and ecosystem services enlists both quantitative and qualitative
methods. Formal surveys and questionnaires typically rely on standardized descriptions of alternative objects/states (e.g., alternative
environmental conditions, management policies, socially-relevant outcomes), with respondents recording explicit choices, rankings, or ratings

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that are analyzed to develop appropriate quantitative metrics (e.g., preference, importance, or acceptance indices). Individual narrative
interview methods typically employ less restrictive representations of options, are frequently directed at specific local cases that are familiar to
respondents, and collect open narrative responses that are subjected to more or less rigorous qualitative analyses. These methods have often
been used to support the design and pre-testing of subsequent quantitative surveys, but they are increasingly being offered as stand-alone
assessments. In addition to the more established methods, some emerging methods base assessments on more direct observations of behaviors
in the environments at issue. Behavioral observation and behavior trace methods have been developed and evaluated, especially in the context
of the assessment of recreation and tourism values (e.g., Daniel & Gimblett 2000, Gimblett, et al. 2001). Computer simulation ("virtual
reality") and interactive game methods are also being developed, but have mostly been applied in research settings (Bishop, et al. 2001a,
2001b). These emerging methods may not yet be sufficiently proven for application in EPA policy-making contexts, but they do show
considerable promise for applications in circumstances where the validity of verbal expressions of preferences and concerns in response to
described hypothetical conditions may be suspect. They will only be briefly described in this section and are offered primarily as potential
targets for future research and development.

Brief description of the Methods

Surveys Including Attitude Survey Questions

Attitude surveys encompass a broad range of methods for systematically asking people questions and recording and analyzing their
answers (e.g., Dillman 1991, Krosnick 1999, Schaeffer and Presser 2003, Appendix C to this report). Questions may assess knowledge, beliefs,
desires, or behavioral intentions about a virtually unlimited range of objects, processes, or states of the person, society, or the world. Multiple
questions/issues are typically presented and responses are reported as choices (among two or more options), rankings, or ratings. The most
popular survey formats have involved face-to-face, mail, or telephone contacts with individually sampled respondents. Web/Internet media are
increasingly being used and are rapidly becoming more sophisticated, but representative sampling issues require special attention. Open-ended
response formats are less often used, and may pose special problems for quantitative analysis.

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Social-psychological surveys have been extensively used to assess preferences, attitudes, importance, and acceptability of presented
policies, actions, outcomes or the expected personal or social consequences thereof (see the lists in Appendix C). An example is the extensive
national survey conducted to support the USDA Forest Service GIPRA process (Sheilds, et al. 2002), which is discussed in Text Box 11.
Multiple value dimensions (e.g., utilitarian, aesthetic, ethical) may be addressed within and between different surveys, and surveys may specify
individual/personal, household/family or social/civic constituencies. The indices produced by application of appropriate quantitative analyses
of recorded responses usually claim to be only ordinal (ranks) or roughly interval scale, relative measures of differences in assessed values
among offered alternatives. Moreover, expressed preferences or other value judgments are assumed to be at least in part created in the context
of the survey (Schaeffer and Presser 2003). Thus, generalization of obtained values measures (e.g., "values transfer") beyond the objects
specifically assessed within a given survey must be approached with caution.

Text Box 11: National Telephone Survey

A nationwide telephone survey was conducted to provide support to the USDA Forest Service Strategic Plan for 2000 required by the
Government Performance and Results Act. The survey randomly sampled over 7000 U.S. citizens to determine held values relevant to public
lands, preferred objectives for management of public forests and grasslands, beliefs about what the role of the Forest Service should be with
regard to these objectives, and public attitudes about the job the Forest Service is doing toward fulfilling the desired objectives. The items for
this "VOBA" survey were developed and pre-tested through more than 80 focus groups conducted across the country. Individual survey
respondents were presented with only a subset of the 115 items/questions developed for the survey. Each respondent assigned ratings to the
items presented on five-point scales, with objective statements (30 items) rated on an importance scale, beliefs (30 items) and values (25 items)
rated on a disagree-agree scale and attitudes (25 items) rated on an unfavorable-favorable scale.

Some example items from the survey and their mean ratings over the full national sample are presented in Table 7 below. Items are selected for
potential relevance to C-VPESS interests and they are grouped to display the observed discrimination in responses. Many of the same items
were rephrased and repeated in several of the values, objectives, beliefs, and attitudes categories (across but not within respondents). Only the
values and objectives category formats and mean ratings (agreement and importance, respectively) are presented here, as the beliefs and
attitudes items were specific to the Forest Service. Some items may be reversed from the original presentation so that higher means always
indicate higher agreement/importance ratings.

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T.ihk* r.Mimpk' I loins I'nnn Sui\o> Supporting I Sl).\ I'oivxl Sitxici* Sliiilo'^ic I'liiii I'm- 2000 minimi In I lit- (.omiiiiikiiI PciloiniiiiHo ;iikI Kosn lis Acl



\ ill lies

()hjecli\es

Item Kxiiinplos

Mean
. 1 greement

Mean Importance







Wildlife, plants, and humans haw equal rights In li\e and jjrow

4.28



future generations should he as important as the current one in the decisions

about public lands

4.52



We should actively harvest more trees to meet the needs of a much larger human

population

: ss



I he decision to develop resources should he made moslh on economic rounds

2.92









Protecting ecos\stems and wildlife habitat



4.58

Conscrv niij and protecting forests and grasslands that are the source of our water
resources, such as streams, lakes, and watershed areas



4.73

fxpanduiij access for motorized olf-hiijhwa\ \chicles on forests and grasslands
(lor example, snow mobilmtj or 4-wlieel driving)



2.41







Designating more wilderness areas on public land that stops access for
development and motorized uses



3.84

Developing new paved roads on forests and grasslands for access for cars and
recreational vehicles



2.62







1 am ijlad there are National forests even if 1 never see them

4 h(i



1 would be willing to pa\ five dollars more each time 1 use public lands for
recreational purposes (for example, hiking, camping, hunting)

s 4w



Individual item standard deviations ranged from 0.75 to 1.50. Sample sizes were not reported per item, hut would he large (at least hundreds of respondents each) so thai
standard errors of the reported means would he very small.

Respondents also answered a mini Iter of demographic questions and pro\ ided inlbmialion about their use of public forests and their knowledge
of and association with the forest ser\ice These items were used to identify se\eral sub-groups that produced different patterns of response to
the items in the sur\ey I 'or example, the authors report. "Metropolitan residents in both the I last and W est see the objective of protecting

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ecosystems and wildlife habitat as more important than do those in non-metropolitan areas. Within non-metropolitan areas, those in the East are
more in favor of such programs than are westerners." p 11.

Similar surveys could obviously be designed to address items relevant to EPA efforts to protect ecosystems and services. The example Forest
Service survey was targeted on broad national strategic goals and issues, but surveys may be even more effective in assessing beliefs,
preferences, and attitudes about more specific management alternatives and outcomes. In some cases, where the relevant dimensions of
outcomes may be subtle and difficult to describe in words, visualizations and other perceptual representations may be more effective in eliciting
public preferences (as illustrated in Text Box 12, Perceptual Surveys, later in this section).

Shields, D. J., Martin, I.M., Martin, W.E., Haefele, M.A. (2002) Survey results of the American public's values, objectives, beliefs, and
attitudes regarding forests and grasslands: A technical document supporting the 2000 USDA Forest Service RPA Assessment. General
Technical Report, RMRS-GTR-95. Fort Collins, CO: U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station. 111 p

Surveys have become ubiquitous in modern society, with uses ranging from assessments of diners' satisfaction with the service at a
restaurant to citizens' support for major national policies (Dillman 2002). Surveys are now frequently directed by computer programs that can
select and order questions individually for each respondent, sometimes based on responses to prior questions. Increasingly, surveys are fully
implemented by computer, allowing the respondent to control (with more or less restriction) the pace of questions and to record their responses
directly into a computer database by key presses, clicks, or voice commands (Tourangeau 2004). Internet-based methods, whose use is
increasing, offer extended possibilities for contacting respondents, presenting questions, and recording responses. However, Web surveys may
raise representative-sampling and other issues that require special attention (e.g., Couper 2001, Tourangeau 2004, and Appendix C to this
report).

Variations on survey research methods that may be especially appropriate for assessments of ecosystems and services include perceptual
and conjoint representations of assessment targets. In perceptual surveys, assessment targets (e.g., existing environmental conditions and/or
projected policy outcomes) are represented by photographs, videos, computer visualizations, audio recordings, or even chemical samples
representing different smells. As for verbal surveys, responses are typically choices, rankings, or ratings of the offered alternatives. Perceptual
surveys may be seen as extensions of traditional psychophysical research methods that have long been applied to assess qualities and

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preferences for foods and other products that are difficult or impossible to describe effectively with words (Daniel 1990). Relevant examples
include assessments of the visual aesthetic effects of alternative forest management policies in the northwestern United States (Ribe, et al. 2002,
Ribe 2006), of in-stream flow levels on scenic and recreational values (e.g., Heatherington, et al. 1993), of visibility-reducing air pollution on
visitor experience in National Parks (e.g., Malm, et al.1981), and assessment of the annoyance produced by aircraft over-flight noise in the
Grand Canyon (Mace, et al.1999). An illustration of perceptual survey methods based on Ribe, et al. (2002) is presented in Text Box 12.

Text Box 12: Perceptual Surveys

A study by Ribe et al. (2002) provides a good illustration of a perceptual survey employing computer visualization technology. The focus of
this study was on the aesthetic effects of the shift to more ecologically motivated forest management in the Northwest United States. The
survey sought to determine how the Northwest Forest Plan (NFP, arising out of the spotted owl controversy) would affect the perceived scenic
beauty of affected landscapes in public forests. Another study objective was to investigate the possible contributions of landscape design
principles contained in the U.S. Forest Service Scenery Management System for shaping NFP harvest prescriptions to provide better aesthetic
results. The description here will focus only on the visualization and perceptual survey components, and how these methods were used to attain
quantitative measures of the aesthetic affects of shifting the emphasis in forest management from economic to ecological goals.

The basic strategy of this assessment was to first select a representative set of forest areas where the NFP prescribed changes to forest
management. From within these areas, 15 forest landscape scenes ("vistas") were selected to represent a range of forest conditions consistent
with pre-NFP management practices. Geographic information system technology was used to delineate and to create 3-D terrain models and
detailed maps of the existing vegetation cover in the visible area of each scene (from a designated viewpoint). GIS perspective view techniques
were used to create a "virtual photograph" of the scene so that color-coded vegetation features (e.g., existing forest, clearcuts of various sizes,
and stages of re-growth) could be accurately located within the view. An actual photograph was also taken from the viewpoint and was
compared with the virtual view to assure accuracy. Forest harvest and growth models and expert judgments of trained foresters working in the
study area were then combined to develop detailed forest management plans for the area within each selected view, following NFP
prescriptions, and to project changes in forest vegetation (removal and re-growth) over 20 years following the implementation of the NFP. A
virtual photograph was again created to represent the projected changes in the visible landscape. Finally, digital montage methods were used to
map appropriate video textures (e.g., five-year re-grown clearcut, undisturbed mature forest, etc.) onto the scene to create a biologically
accurate but photographically realistic visualization of the future forest conditions. Figure 9 below illustrates some of the key steps in this
visualization process.

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Figure 9: Visualization of Forest Conditions and Actual Photos from Ribe et al. (2002)

2

3

4

5

6

7

The visualizations of future forest conditions and the actual photos for each of 15 selected study scenes were rendered to color slides. Study
slides were intermixed with 90 additional scenes representing a wide range of forest conditions in the area and presented in a perceptual survey.
The 608 respondents were sampled (not randomly) from 31 stakeholder groups in the Cascade region affected by the NFP. Respondents
recorded their judgments of the scenic beauty of each scene independently on an 11-point scale ranging from "extremely ugly" (-5) to
"extremely beautiful" (+5). Because all scenes were rated by the same groups of respondents in the same context, simple mean ratings were
judged an appropriate index of the relative scenic beauty of the scenes. Because the study scenes were specifically selected to represent

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particular forest management-by-\ iew parameters (not random samples) comparisons were restricted to the pre-NTI' \ ersus post-NI-l' pairs for
the same base scenes

The mean differences between pre- and post-\l P pairs for the I 5 forest scenes ranged from -3.05 (fa\ oring the pre-\l P prescription) for a
close-lip \iew of a recent har\est to 2 3l> la\orinu the \l P prescription in a larger scale \ista with numerous har\est sites in the \ i si hi e area.
I'or six of the eight scenes selected to ha\ e large to medium-sized \ iew areas, scenic beauty ratings were significantly higher for the post-\l P
scene Regression analyses determined that the key. ohjecli\ely measured \ariables affecting scenic beauty differences between pairs of scenes
were the percent of the \ i si hie area co\ ered by fresh, high-contrast clearcuts in the middle distance and in the far distance of the \ iew (both
with negati\e coefficients)

The Ml' management prescriptions were primarily dri\en by ecological considerations, but in the most conspicuous cases (the larger \iews)
these ecological prescriptions also produced significant impro\ ements in scenic beauty as percei\ ed by the most likely \ isitors to and \ iewers
of those and similar sites While this study did not directly address the question, a similar perceptual sur\ey. along with standard forest
\egetation co\er and har\est data could be used to measure and map trade-offs among economic, ecological and aesthetic \ allies for forest
management alternate es (including MP and other approaches) based on a systematic sample of \ iew points scenes across a landscape of
interest Such trade-off assessments and regression-based models can be used by forest planners to dc\elop detailed har\est prescriptions and
schedules for specific sites allowing \l P ecological guidelines to be met while maintaining or enhancing economic and aesthetic goals for the
public landscape. In LPA contexts, similar perceptual sur\ey methods might be applied to assess aesthetic and other \isual impacts at
contaminated sites, and to assess the relati\e merits of restoration and reuse options

Ribe. R (i.. Armstrong. I- T .tK; (iobster. I* 11 (2()iP) Scenic \ istas and the changing policy landscape \ isualixing and testing the role of
\ isnal resources in ecosystem management. Landscape Journal. 21 42 (•>(•>

Surveys most often present the individual attributes of assessment targets separately. For example, a survey to assess the effects of a
proposed environmental policy might present separate questions to determine respondent's judgments about effects on air quality, water
quality, and local employment. Conjoint survey questions (e.g., Adamowicz, et al.1998, Boxall, et al. 1996) instead present options as
multidimensional composites or scenarios presenting integrated/conjoined combinations of different attributes (e.g., different levels of air
quality, water quality, and local employment). Combinations generally reflect actual or projected variations in the attributes (e.g., different
levels of air and water quality and local employment opportunities). In the more sophisticated conjoint surveys, the particular combinations of

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attributes represented are specified by an experimental design that allows estimates of the separate and interacting effects of component
attributes (Louviere 1988). Multiple regression (or similar) analyses are used to estimate the relative contributions of individual components
(attributes) to the expressed preferences (or other judgments) for the conjoint alternatives.

Conjoint survey questions can provide relatively direct estimates of the value trade-offs people make when choosing among outcomes
composed of multiple attributes that naturally covary and whose values potentially conflict and compete. When at least one of the attributes that
forms the conjoint alternatives is (or can be) valued in monetary terms, the regression equation based on expressed preferences among the
conjoint alternatives can be translated so that coefficients for all attributes are expressed as monetary values (see the following Appendix B
Section on Economic Methods). An illustration of conjoint survey methods is presented in Text Box 13: Conjoint Surveys.

Text Box 13: Conjoint Surveys

Conjoint methods may be especially well-suited for gauging public preferences across sets of complex multi-dimensional alternatives, such as
alternative EPA regulations or management options for ecosystems/services protection. Respondents choose among (or rank or rate) multi-
dimensional "conjoint" alternatives that present specific packages of desired and less-desired attributes. Analyses of the patterns of preferences
values (e.g., probability or percent choice or mean rating) among the conjoint alternatives can be used to estimate the contribution (e.g.,
regression coefficients) of each of the separate attributes.

Chattopadhyay, Braden, and Patunru (2005) used a conjoint survey method to assess the effects on residents' home preferences of various
cleanup options for the Waukegan Harbor Superfund site in Wisconsin. This study also employed and compared results of a hedonic pricing
method (see the following Appendix B Section, Economic Methods, for a description), but the monetary estimates of willingness to pay for the
cleanup options evaluated were based on stated preferences in a conjoint survey, which is the subject of this illustration. Adjustments for
differences in respondents' incomes, annual costs for current housing and for the hypothetical housing options offered (based on real estate
market data) and a number of composite and interaction terms involving economic variables were introduced to conform to assumptions of
relevant economic theory and practices. However, the basic data are simply respondent's choices (expressed preferences) among alternative
hypothetical conjunctions of housing and environmental-condition attributes, so the core features of the study nicely illustrate an application of
a conjoint choice survey that could just as easily, or more easily, be used to obtain an interval scale measure of the effects of cleanup options on
housing preferences.

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1	lousing market data lor 47. |oo iransaclions (Iiwr>-2<)()I) for Waukegan and 12 similar nearby cities along with focus group sessions with
homeowners were used to determine the six housing en\ ironmental attributes that were conjoined to describe the hypothetical housing options
and to describe the respondents" own current home en\ ironment I lousinu attributes were lot size, house size, and house price. l-n\ ironmental
attributes were elementary school class size, public areas near the harbor, anil extent <>J changes proposal in the harbor-area pollution l-ach
of the (¦> housinu en\ ironmental attributes was represented by four le\ els. so that in principle there could be 4" 4 distinct conjoint options.
A fractional factorial experimental design (with a "lbld-o\ei ' to allow estimation of two-way interaction terms) was used to determine the M x

2	I2S conjoint options that were actually assessed in the sur\ey The details and rationale for this complex ilesiun is beyond the scope of this
illustration, but the key point is that the selected alternate es allow for statistical estimates of the separate effects of each of the

housinu en\ironmental attributes on o\erall preferences (or o\ oral I w-t-p estimates in the present study) across all of the options All I2S
selected options were assessed in the study, but each of the l>54 respondents (from 233^ sur\eys mailed to the 13 targeted communities) only
responded to a random subset of !(•> options

In a typical conjoint choice study, respondents would see pairs of the conjoint house en\ironment options and be required to choose between
them Challopadhyay el al instead chose to reduce the length and complexity of the task by comparing each hypothetical alternate e to a
standard the respondent's current home en\ ironmental conditions The difference on each of the six attributes between the current home and
each hypothetical option was expressed as a percentage l or example, the house size attribute could be I 5"o smaller, unchanged. I 5"o larger or
25"o larger than the respondent's current home, and the harbor area environmental condition could be additional pollution, no change (from
current conditions), partial cleanup, or full cleanup A facsimile of an illustrati\e choice question in the sur\ey is presented in Table X below

I ;il)k* S: I'iicsimik' ol' Illiislr;ili\o Choice Oiioslions I'mm < h;iIlopudh>n> el ;il. (2005)

1 Ionic 1 Imagine your home modified to 111 this description



1 .ol size

1 louse
size

School class
size

Public natural areas in
harbor area

1 larbor area

en\ ironmental condition

1 louse price

Compared to your
current home

Smaller
by 1 5" ci

Smaller
by 1 5" (i

Smaller by
2 students

Smaller by 2<)"„

Additional pollution

1 .ess expensi\e by l<>"(>

The core data for llic conjoint choice study is the obser\ eil probability of choice for each of the 12S hypothetical house en\ ironment options
o\er the current home These probabilities can be used to deri\c more sophisticated i|iiantitati\e \alue scales, but basically the worst options

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1	(least preferred) would be chosen less often and the best would be chosen more often. In conjoint studies, choices for the hypothetical multi-

2	attribute options is usually of less interest than are the estimates of the contributions of the respective house/environment attributes to those

3	expressed preferences. There are numerous methods for attaining these estimates, most based on multiple regression analyses of one kind or

4	another. In the Chattopadhyay, et al. study, a multinomial/conditional logit model was used. The details of this analysis are not relevant to this

5	illustration, but the basic outcome of such a conjoint choice study can adequately be portrayed as a regression equation of the following form:

6

7	Pi = Wi(Ali) + W2(A2 i) + W3(A3 i) + W4(A4 i) + W5(A5 i) + W6(A6 i)

8

where

Pi is probability of choice (versus current home) of conjoint alternative i
wi is the regression coefficient for house/environment attribute 1 (e.g., lot size)

Al; is the level for attribute 1 for alternative i (e.g., 15% smaller)
and so on for each of the other 5 house/environment attributes.

Chattopadhyay, et al. scaled the weights in a much more complex equation (including derived economic terms and interactions) to attain
monetary benefit estimates on the basis of which they offered conclusions such as:

.. .the significant coefficient for the interaction variable full*highhic indicates that high-income residents prefer full cleanup more than
other categories, while the insignificant coefficients on ciddpol*highinc and pcirt*highinc indicate that high-income residents are no
different from others (income levels) with respect to their dislike for additional pollution and their preference for partial cleanup, p 367

The authors went on to estimate aggregate monetary benefits of partial and full clean up of the Waukegan Harbor Superfund site ($249 million
and $535 million, respectively). The validity of these monetary estimates, of course, depends upon a complex set of assumptions required by
general economic theory and by specific features of the present study. These assumptions would not be required for the more basic analysis of
expressed preferences suggested in this illustration. The attribute weights (regression coefficients) in the suggested simple preference equation
could, however, safely be interpreted as relative (interval scale) measures of the trade-offs the sampled respondents made between the offered
changes in harbor environment cleanup (from additional pollution to full cleanup) and the other house/environmental attributes represented by
the options in the study.

Once determined, the preference-based regression equation could also be used to estimate preferences for new policy alternatives based on their
respective projected changes in environmental conditions, so long as those options fit sufficiently within the range of the attributes and levels

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1	assessed and the constraints imposed by the context of the survey in which the house/environmental condition options were offered and judged.

2	Optimization or less formal heuristics might be applied to create additional policy options for consideration and/or for direct evaluation in

3	subsequent conjoint surveys.

4

5	Chattopadhyay, S., Braden, J. B., & Patunru A. (2005) Benefits of hazardous waste cleanup: new evidence from survey- and market-based

6	property value approaches. Contemporary Economic Policy, 23, 3: 357-375.

7

8	Individual Narratives

9	Researchers using individual narrative methods contact individual respondents, who participate alone, without interaction or discussion

10	with experts, facilitators, or other respondents. Individuals nominally representing possible stakeholder perspectives are contacted and asked to

11	comment on relatively broadly defined topics with relatively little direction from the interviewer/assessor (e.g., Brandenburg & Carroll 1995).

12	Respondents are not typically selected by a random, probability sampling process. Instead, particular individuals are specifically targeted

13	because of their known or assumed nominal group membership or personal relationship to the problem/policy/outcome at issue. The sample

14	may be extended by having prior respondents refer others, as in the "snowball" technique. The number of individuals to be included is quite

15	variable, and in a relatively few cases has been determined by some formal process based on a rolling analysis of collected narratives (e.g.,

16	using a criterion of diminishing new perspectives/positions being discovered). Collected narratives are subjected to more or less rigorous

17	qualitative analyses, (essentially similar to the analysis of focus group responses, see Appendix B section on Focus Groups) to explore and

18	articulate the breadth and depth of expressed understandings and concerns relevant to the assessment target. Included in this category are

19	various ethnographic methods and mental modeling procedures.

20	A mental models approach can inform debate about the best ways to elicit values, and how people use and understand different

21	qualitative and quantitative expressions of value, response scales and response modes. People use their prior (pre-existing) mental models to

22	interpret survey questions and other preference-elicitation probes. People make inferences not only about texts in surveys, but also about values

23	and risks in the actual environment and hence their mental models and mental representations of causal processes underlie all decisions. Mental

24	models methods aim at eliciting people's understanding of causal processes associated with the events, processes, and actions that are projected

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to result from specific decisions. As applied to understanding hazardous processes, the method has been used to characterize people's
understanding of how risks arise and can be mitigated, and entails a mixture of decision modeling, semi-structured interviews (ethnographic in
nature), survey research, comparisons between these, and both qualitative and quantitative modeling of the results. To date, this research has
focused more on enabling and informing risk reduction, rather than motivating or understanding preferences and trade-offs per se.

Mental models research would be an appropriate precursor (i.e., formative analysis) to any formal survey or preference elicitation
method, to improve the validity and reliability of the method. Values are typically expressed qualitatively, sometimes in ordinal terms (e.g.,
lexicographic scales or comparative statements) and sometimes using quantitative scales. The approach is designed to explore the conceptual
landscape for risks and benefits, including underlying causal beliefs, specific terminology/wording, and the scope and focus of mental models
in the decision domain of interest. The approach is principally qualitative, designed to elicit how an individual conceptualizes and categorizes a
process, such as protecting an ecological service, and how that individual would make inferences about and decisions to influence that process.

Emerging Methods

The assessment methods described in this section are relatively new and untested. They are characterized by more direct observation of
responses to policies, outcomes and consequences in situ, avoiding problems of relying on hypothetical responses to described conditions. In
that context, these methods parallel the revealed preference methods used in economic value assessments (Appendix B, Economic Methods).
Observed environmental behavior is often not consistent with what people say they would do in the specified circumstances (Cole and Daniel
2004) and people are often incorrect at identifying, or are unaware of the environmental factors that affect their behavior (e.g., Nesbitt and
Wilson 1977, Wilson 2002). In the context of ecosystems and services, behavioral observation methods monitor the activities of people in a
particular environmental context and observe changes in behavior as relevant conditions change over time within a site or over sites with
differing characteristics. Behavior trace methods are based on indirect evidence of people's behavior in specific environmental contexts. For
example, the number of visitors to recreation sites might be estimated by counting the number of autos parked at access points, by the number
of passers-by recorded by automated trail counters, by the number of fire rings in dispersed camping areas or by the amount of trampling and

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disturbance of vegetation along trails and at destination points. Direct observations or traces of visitors' activities can be correlated
geographically with relevant environmental/ecological conditions or monitored over time as changes in conditions occur at the same sites,
revealing the effects of these changes on environmental preferences and reactions (e.g., Gimblett, et al. 2001, Wang, et al. 2001, Zacharias
2006).

These methods do not seem to have been applied in the context of assessments of the effects of changes in ecosystems and services.
However, changes in human use of rivers, lakes, and estuaries are often important indicators of the need for and the value of EPA interventions
to protect water quality and associated aquatic systems, and the travel cost methods employed by economists in these contexts is fundamentally
similar. Behavioral observation and trace methods might be effectively employed to attain quantitative measures of human use levels that
could be used in conjunction with economic measures or as separate measures to be correlated with changes in ecological conditions. Numbers
and durations of users, their geographic distribution and the activities that they engage in might be correlated with relevant bio-physical
measures of ecological conditions to develop useful assessments of the effects of ecological degradation or the effectiveness of ecological
protection efforts.

Interactive environmental simulation systems provide means to overcome some of the limitations and difficulties of conducting direct
behavioral observations or interpreting behavior traces. Direct observation methods are necessarily limited to existing conditions and are
potentially confounded by uncontrolled or unrecognized irrelevant variables. Most policy decisions hinge on people's responses to specific
changes to not-yet-existing, projected environmental conditions. Rapidly advancing computer technology has enabled effective and
economical simulation of complex dynamic environments at high levels of realism (e.g., Bishop and Rohrmann 2003, Bishop, et al. 2001a,
2001b). The emphasis has been on visual presentations, but the technology can readily include auditory features and in some systems tactile,
proprioceptive, olfactory, and other senses can also be effectively simulated to achieve very compelling, immersive environmental experiences.
Moreover, expanding response options, ranging from the computer mouse to video-game controllers to gloves to full-body movement enable
increasingly natural interactions with simulated environments. In the context of assessing the effects of changes in ecosystems and services,
interactive computer simulation systems offer the opportunity to conduct virtual in situ experiments to determine how persons respond to

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specific investigator-controlled changes in environmental conditions. Thus the effects of manipulated conditions on environmental preferences
and other reactions can be revealed in a context closely approximating "real world" circumstances.

Interactive computer simulation systems may be viewed as games, in which human respondents attempt to (virtually) navigate through
and perhaps alter virtual environments to accomplish desired goals. There may be no particular outcome that can be defined as "winning" such
a game, but the behavior of the player and the outcome on which s/he settles can reveal the values that motivate and guide the player's
responses. Interactive games can be informative in this regard, even if they are played in substantially less than virtual environments. Indeed,
more limited and/or more abstract games may have important advantages in some circumstances. For example, it may not be possible to
project the explicit and detailed outcomes of a proposed policy that are required for a realistic environmental simulation, and the specific
implications of particular responses to changing environmental conditions may not be known. In many situations only changes in some
particular ecological component may be known and relevant (e.g., a reduction in a particular contaminant or an increase in survival rates of a
particular wildlife or plant species). Still, a game-like context may be an effective and engaging way to communicate with public audiences
about what outcomes they would prefer, and what policies are required to achieve those outcomes. A major advantage of games over surveys,
for example, is the opportunity for respondents to learn through experience about how the ecosystem of interest responds to various policies or
policy aspects and to progressively modify their expressed policy preferences to converge on some acceptable balance among desired and
undesired outcomes.

Relation of Methods to the C-VPESS Expanded and Integrated Assessment Framework

Survey and individual narrative methods have useful roles to play throughout the valuation process envisioned by C-VPESS. For
example, representative surveys and selected individual interviews could contribute to initial problem formulation by identifying ecological
services and impacts that most concern citizens and/or identified stakeholders, as well as by uncovering assumptions, beliefs, and values that
underlie that concern. Similarities and differences in assessed concerns, attitudes, and beliefs toward proposed policies among different
segments of the public can also be identified and articulated. Once relevant ecological endpoints have been identified, surveys could be very

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1	useful for determining the personal and social consequences of those outcomes, and for exploring public understanding of the links between

2	chains of ecological effects and the policy options under consideration (Box 3 in Figure 2, reproduced below). Given a set of potential policy

3	options, with their respective ecological endpoints (from Box 4 in Figure 2, surveys could be used to assess relative public preferences [and/or

4	other judgments, such as importance or acceptability]) for those options (Box 4 in Figure 2). Quantitative indices of public/stakeholder

5	preferences (or judgments of importance or acceptability) from surveys could be combined with bio-ecological and economic/monetary

6	measures of the value of the same alternatives to provide cross validation for all measures and to strengthen the foundation for policy decisions.

7	Surveys may be especially useful when the values at issue are difficult to express or to conceive in monetary terms or where monetary

8	expressions/valuations are viewed as ethically inappropriate. In those cases survey questions could provide reliable quantitative measures of

9	public preferences among the policy alternatives or ecological endpoints that are under consideration, improving the basis for Agency decision
10	making.

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Attitude survey questions could make an additional contribution after Box 5 in the C-VPESS model. The values of ecosystems/services
coming out of Box 5 must inevitably be represented by multiple economic/monetary, bio-ecological and social-psychological indicators. EPA
administrators can be left with the difficult task of integrating these diverse and potentially conflicting measures, along with legal, budgetary
and other constraints to make and rationalize policy decisions. Properly structured attitude survey questions, perhaps including material to
inform respondents about relevant ecological and social effects and other considerations affecting the policy/decision at issue, could effectively

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involve citizen stakeholders in this value integration and trade-off process, providing an additional relevant input to the policy decision, and
adding to the political validity and social acceptability of the final action.

Individual narrative methods, such as the mental models method, would be most appropriate and most useful at the earliest and latest
stages of the decision making process. While individual interview methods do not generally provide quantitative assessments for alternative
policies or outcomes, they can make important contributions to improving the design, development and pre-testing of more formal surveys that
can provide reliable and valid quantitative assessments of public concerns and values. Mental models methods are appropriate for use in all
identification stages (ecological modeling; what matters; ecological impacts that matter), with the possible exception of identifying EPA's
objective(s). Genuine probing interactions with individuals or groups representing key stakeholders and including divergent views and
concerns should be a central part of problem definition and identification of significant ecological and associated social effects components of
the process. Such interactions with key stakeholders and with citizens could also inform the values integration and negotiation in the final
decision process and guide and pre-test the communication of that decision.

Status of Methods

Survey questions measuring social-psychological constructs are the oldest and most frequently used methods for determining public
beliefs, concerns, and preferences. Survey questions have been used and continue to be used effectively by all levels of government to measure
citizen desires, concerns, and preferences. Economists have lately adapted survey methods to measure stated willingness to pay for non-market
goods and services, and surveys are often relied upon to collect the data needed to exercise other economic valuation efforts, such as travel cost
and hedonic pricing methods (see Appendix B, Economic Methods). Environmental management agencies have made use of surveys, either
directly or indirectly, in setting policy and in making and monitoring the effects of management decisions (e.g., Shields, et al. 2002, illustrated
in Text Box 11 and the many surveys listed in Appendix C to this report).

It is not clear the extent to which individual narrative interviews are systematically used in EPA policy making, nor do the OMB and
other guidelines clearly specify the criteria for using these methods. While no specific evidence has been found either way, it seems reasonable

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to assume that individual narrative interviews have not been important components of formal EPA decision making processes. Certainly the
qualitative nature of the information provided by both focus groups and individual interviews, and the general disinterest in representative
sampling makes them poor candidates for formal policy evaluation exercises, but that does not preclude their having a role in earlier stages of
the decision making process as envisioned by the C-VPESS. Mental models research could in theory be applied as a first step to investigate
either "means" or "ends" values. This method would be an appropriate precursor (i.e., formative analysis) to any formal survey or preference
elicitation method, to improve the validity and reliability of the method.

Limitations

The largest barriers to greater use of survey methods in ecosystems and services valuation and decision making by the EPA are
institutional. First, while the EPA seems to have embraced economic surveys (e.g., CVM, or at least "transfers" from prior CVM surveys) as a
valuation method, there is a noticeable reluctance to use the larger class of systematic surveys using attitude, preference, and intention
questions, relative to the practices of other federal agencies with similar environmental protection mandates and valuation needs. This
predisposition may in part be due to specific legal requirements for formal monetary benefit-cost analyses (which also apply to other agencies),
but none of the currently applicable laws preclude using a fuller range of value measures and methods, and the most prominent laws and guides
explicitly urge a broadly based evaluation effort not limited to monetary measures. Aside from this agency-level barrier, survey methods in
general are discouraged by federal rules implementing the Paperwork Reduction Act. Over the past several decades it has been difficult for
federal agencies to attain required clearances (e.g., from the OMB) for surveying the public in a manner and in a time frame that effectively
addresses policy evaluation needs. This institutional barrier is formidable, and the proliferation of surveys and pseudo-surveys for commercial
and political purposes has dampened citizen's willingness to participate, but many significant surveys continue to be conducted by a number of
government agencies (see Appendix C for further discussion).

When used, survey questions have proven effective for measuring public knowledge, beliefs, attitudes, and intentions. However,
especially in the context of the complex processes of selecting alternative policies and actions to protect ecosystems and services it is important

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to recognize that the responding public may not a priori have a great deal of information or knowledge about the issues or policies about which
they are asked. First, limitations on length and complexity of content (especially for telephone surveys) make it unlikely that the full
complexity, including uncertainties of policies and their outcomes can be effectively communicated to respondents within the survey. Second,
the general public is unlikely to have the breadth and depth of ecological knowledge that is often required to understand and evaluate a given
policy, its bio-physical outcomes or the implications of outcomes for the respondent or for society more generally. Finally, even when the
respondent fully understands these aspects of a proposed policy he/she may still be uncertain (or incorrect) about his/her projection of how well
(or badly) the respondent will feel about the outcomes/implications when they are actually encountered (Wilson, et al. 1989). Some approaches
to addressing these problems in surveys are presented and discussed in Appendix C to this report.

The technical issues that have been of the greatest concern to users of survey information, to quality control agents (e.g., OMB), and to
survey researchers have been associated with the sampling of respondents. The results of a survey are typically intended to be generalized to
some specified population (e.g., adult citizens of the United States) that includes many members that will not be included in the sample of
individuals who actually respond to the survey (the respondents). The integrity of generalizations to the population of interest is assured if the
respondents are a formal representative sample ("probability sample") of the population. However, recent research shows that departures from
strict sampling rules, such as the loss of intended participants by non-response or failed contacts, may not have as strong an effect on the
representativeness of survey outcomes as some have thought. More difficult and potentially more potent errors are in survey design, including
the crafting, selection, and ordering of questions/items to be included in the survey, the form of the response options offered (e.g., the type of
ratings scales), and uncontrolled events that occur during the time of survey implementation (see Krosnick 1999 and Appendix C to this report).

Social-psychological surveys do not meet the requirements of economic cost-benefit or cost-effectiveness analyses because they do not
achieve a unidimensional, transituational measure of value. That is, the scale values computed for the ecosystem and service options addressed
in a survey cannot be directly compared to (may not be commensurate with) values for extra-survey options, or to values and costs in other
domains of the respondents' lives. It is arguable whether any value assessment method fully meets this requirement. However, given a feasible
set of alternative regulatory/protection actions and outcomes in a specified environmental-social context, surveys of public attitudes,

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preferences, and intentions would be appropriate for quantitatively measuring public preferences among offered sets of policy/outcome options,
for estimating the relative importance to people of the multiple attributes of those policies and outcomes, and for gauging the acceptability of
management means for achieving them. Properly designed conjoint methods may be especially well-suited for gauging public preferences
across sets of complex multi-dimensional alternatives, such as will likely be involved in many EPA regulations and actions for
ecosystems/services protection.

In practical use, the human resources required to implement surveys range from a sufficient cadre of technically competent survey
designers and analysts to temporary hourly wage employees to perform the mailing, phoning, or interviewing tasks. Material needs may be
very low ("paper and pencils") or quite high, as when sophisticated computer simulations/visualizations or interactive response formats are
employed. Face-to-face surveys, where trained interviewers are required and participant-contact costs may be high, are generally the most
expensive, but costs for mail, telephone and/or computer resources can also be significant in large surveys using those formats. All of these
costs are usually quite low relative to the physical, biological and/or ecological science and field study required to create adequate projections
and credible characterizations of value-relevant means and outcomes for a suitable range of alternative regulatory or protection actions. In
many ways, the quality of evaluations of ecosystems and ecosystem services protections most depends upon the quality of the relevant
projections and specifications of ecological endpoints and their social consequences. In some cases considerable resources may have to be
devoted to translating targeted ecological outcomes into understandable representations of socially relevant effects. Once these essential factors
have been accomplished, the cost of a systematic public value assessment survey can be comparatively quite small.

Individual interviews can have important and useful roles to play in Agency policy and decision making. However, their emphasis on
qualitative analyses and their typical disregard for representative sampling can make them less useful for formal evaluations or comparisons of
alternative policies and outcomes. These methods can be very useful and important for designing and pre-testing more structured surveys that
do provide quantitative assessments of values for alternative policies and outcomes. Qualitative methods may also contribute to the design of
more effective communications and rationalizations of Agency decisions to stakeholders and to the general public. In mental models research,
values may be expressed qualitatively, sometimes in ordinal terms (e.g., lexicographic or comparative statements), and sometimes using

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quantitative scales. The approach is designed to explore the conceptual landscape for risks and benefits, including underlying causal beliefs,
specific terminology/wording, and the scope and focus of mental models in the decision domain of interest. A mental models approach would
best be used in conjunction with another method in order to obtain quantitative measures of values. The approach is qualitative, designed to
elicit how an individual conceptualizes and categorizes a process, such as protecting an ecological service, and how that individual would make
inferences about that process, as well as any decisions to influence it.

Treatment of Uncertainty

Survey methods specifically address the uncertainty introduced by sampling errors (e.g., representative sampling, non-response),
specification errors (e.g., adequate descriptions or representations of alternatives, clear and understandable response system), and the effects of
a variety of contextual and external factors that may affect (bias) participant responses. Methods for reducing and quantifying the magnitude of
most of these sources of uncertainty and error in surveys are part of the well-documented technology and the accumulated lore of survey
research (e.g., Dillman 1991, Krosnick 1999, Tourangeau 2004, and Appendix C to this report).

Accepted methods are available and are commonly used for calculating confidence intervals or complete probability distributions for
individual survey responses over respondents (e.g., the importance ratings assigned to a particular item). The internal reliability and
cohesiveness of survey responses can be calculated per individual respondent, but more often the focus is on the mean response of
homogeneous groups of respondents. Multiple items are frequently combined, as by cluster or factor analysis, into latent variables (factors)
implied by the inter-correlations among individual-item responses, and there are several conventional statistical indices of the internal
consistency and coherence of those derived factors. More complete analyses calculate and quantitatively assess the internal consistency and
distinctiveness of latent variables, based on the patterns of responses across the multiple respondents, as well as classifying sub-groups of
respondents, based on patterns of individual's responses to the multiple items in the survey.

The detailed results of a survey of a representative sample of a population are unlikely to be fully appreciated by anyone without relevant
training and experience. On the other hand, results can be, and routinely are, simplified for communication to lay audiences. Most people

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would find reports such as "alternative A was preferred over all others offered in the survey by 75% of respondents" to be clear and intuitively
understandable. A table or graph showing mean preference ratings on a 10-point scale for all alternatives evaluated would be clear to many
members of the public, as well as to experts from other scientific and managerial disciplines that are involved in EPA rule and decision making.
Some of the uncertainty associated with these indices (e.g., sampling and measurement error) could be displayed by conventional confidence
intervals or error bars. The potential effects of more complex sources of uncertainty might be revealed by bracketing mean estimates for each
alternative assessed with 25th and 75th percentile estimates derived from sensitivity analyses exercised over the entire biological-social
evaluation system. The most sophisticated communication devices might be based on interactive game systems, where the audience is allowed
to alter input variables and assumptions about functional relations and stochastic events and observe and learn for themselves how these
changes affect projected evaluation outcomes.

Research needs

Issues that should be addressed in future research relevant to social-psychological value assessment methods include:

•	How can structured surveys of public/stakeholder attitudes, preferences, and intentions best be used in EPA policy and decision
making, including how decision makers can and should use the relative quantitative (non-monetary) value indices provided?

•	How can social-psychological value indices best be used to cross-validate estimates of monetary values (e.g., from CBA) and
ecological indices (e.g., biodiversity, energy flow) and strengthen the basis for Agency decisions about alternative
ecosystems/services policies?

•	How, and when in the decision process, can social-psychological, economic, and bio-ecological evaluations of changes in
ecosystems and ecosystems services most effectively be integrated to support Agency policy and decision making?

•	What productive roles can individual interviews and other qualitative methods play in Agency policy and decision making?

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• How might the development of emerging methods (behavior observation, behavior trace, interactive computer simulations, and
games) be shaped to effectively contribute to Agency policy and decision making needs?

References

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scale. Journal of Social Issues, 56, 425-442.

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ECONOMIC METHODS

Overview

Brief Description of Methods

The economic concept of value is based on two fundamental premises of neoclassical welfare economics: that the purpose of
economic activity is to increase the well-being of the individuals in the society, and that individuals are the best judges of how well off they
are in any given situation and of what changes would enhance that well-being.

The concept of value underlying economic valuation methods is based on substitutability, or, more specifically, on the trade-offs
individuals are willing to make for ecological improvements or to avoid ecological degradation. These trade-offs provide an indication of
changes in well-being that result from increases and decreases in goods and services people value. By itself, an ecological change that an
individual values will increase that person's utility. The value or benefit of that change can be defined in two ways. The first is the amount
of another good that the individual is willing to give up to enjoy that change (his "willingness to pay" or WTP). The second is the amount of
compensation that a person would accept in lieu of receiving that change (his "willingness to accept" or WTA). These trade-offs are typically
defined in terms of the amount of money an individual is willing to pay or willing to accept and hence benefits are measured in monetary
terms. In this case, WTP is the amount of money that would make the individual indifferent between paying for and having the improvement
and foregoing the improvement, while keeping the money to spend on other things. Likewise, WTA is the amount of money that would
generate an increase in utility equivalent to that realized from the improvement in the environmental amenity.

However, it is important to note that the concept of benefit does not hinge on the use of monetary units. In principle, benefits could be
defined in terms of changes in any other good or service that the individual would willingly agree to in exchange for the environmental
change (e.g., food). The use of money as the basis for exchange is simply a convenience. In particular, use of a common money metric
allows all benefit measures to be easily aggregated and compared with monetary measures of cost.

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The benefits captured by the concepts of WTP or WTA can be derived not only from goods and services for which there are markets
(e.g., forest products) but also from goods and services for which markets might not exist (such as clean air and clean water). In addition,
they include values derived from use of the environment (e.g., hiking in the woods) as well as those derived from the "existence" of a valued
species or condition. Thus, economic valuation captures values that extend well beyond commercial or market values. However, it does not
capture non-anthropocentric values (e.g., biocentric values) and values based on the deontological concept of intrinsic rights.

All economic measures of value based on willingness to pay are limited by the fact that the maximum amount a person could pay for
anything is constrained by that person's ability to pay, which is indicated by the individual's wealth. Thus the value estimates derived from
economic valuation methods are conditional on the existing distribution of income and prices. As a result, acceptance of these benefit
estimates implies acceptance of the underlying distribution of wealth. One way to incorporate concern for equity in the distribution of well-
being, with roots going back to Bergson (1938), is to weight the measures of economic value or welfare change for each individual by that
person's relative degree of "deservingness"; that is, to attach a higher weight to benefits going to those judged to be more deserving because
of some attribute such as their lower level of income. However, there is no clear way to determine the appropriate weights. In practice,
analysts typically use the value measures derived from the mean individual in the sample that is providing data for the valuation model in use.
If value or willingness to pay is an increasing function of income, the analyst is implicitly underestimating the values of the highest income
individuals and overestimating the values of the lowest income individuals. The result, in a crude qualitative sense at least, is equivalent to
assigning more weight to the values of low income than high income individuals.

The key input for all of the economic methods is data on the choices that people have made or indicate they would make about the
things that contribute to their economic well-being. These choices are made in several contexts. The first is choices about quantities
demanded and supplied in markets at alternative prices, e.g., the amount of commercial fish that are harvested and sold at various prices.
These choices generate demand and supply functions that can be estimated with the information on the amounts purchased at different prices
using statistical (i.e., econometric) methods. Changes in these demand and supply functions in response to changes in the levels of ecosystem
services (e.g., a change in water quality) can be analyzed to obtain market-based estimates of the values of the changes in these services.

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Second, choices can involve the selection of quantities of goods and services (or responses to changes in the availability of goods and
services) that are not sold in markets, such as many ecosystem services. Non-market revealed preference methods can be used to obtain
estimates of the values of changes in these goods and services. Third, hypothetical choices made in response to survey questions can be
analyzed with one of the several stated preference methods for valuation to provide information on trade-offs people would be willing to
make. The specific methods that employ these three different types of choice data to value ecological changes are discussed in more detail in
the following sections.

Key References:

Bergson, A. 1938. A Reformulation of Certain Aspects of Welfare Economics. Quarterly Journal of Economics 52:310-334

Bockstael, Nancy E., and A. Myrick Freeman III. 2005. "Welfare Theory and Valuation," in Karl-Goran Maler and Jeffrey R. Vincent, eds.,

Handbook of Environmental Economics, Amsterdam: Elsevier.

Champ, Patricia A., Kevin J. Boyle, and Thomas C. Brown, eds. 2003. A Primer on Nonmarket Valuation, Dordrecht: Kluwer Academic
Publishers.

Freeman, A. M. III. 2003. The Measurement of Environmental and Resource Values: Theory and Methods. 2nd ed. Washington, DC:
Resources for the Future

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Market-Based Methods
Brief Description of Method

The market-based approaches to economic valuation are used to estimate the economic values of ecosystem services that are an input
into the production of a good or service that can be bought and sold in a market at an observable price. For private goods and services
purchased in competitive markets, the price of a good reflects the valuation of an extra unit of that good or service by the set of participants in
that market. For small changes, market prices can be used as a measure of economic value of each unit of the goods involved. For larger
changes, however, marginal willingness to pay (demand) and marginal cost (supply) are unlikely to remain constant, requiring estimation of
changes in consumer and producer surplus. 44

This approach can be applied in a variety of contexts. For example, wetlands often serve as nurseries for fish species that are
harvested for commercial markets. They are thus an input to commercial fishing, and their services affect the supply and market price of
harvested fish. The economic benefits of protecting wetlands can then be estimated by their contribution to the market value of the output of
the commercial fishery. For relatively small changes, the additional output of the fishery can be valued simply by multiplying the change in
output by the market price of the fish. Similarly, when a river is used as a source of irrigation water for agriculture, both the water quantity
and quality directly contribute to the production of food. The economic benefit of an improvement in either water quantity or quality can be
estimated by its contribution to the market value of food production. Again, for small changes, the market price of the agricultural product
multiplied by the resulting change in output provides a measure of the value of the water quality or quantity change.

Status as a Method

Market-based methods are based on well-established economic principles and econometric practices (Boardman, et al. 2006,
McConnell and Bockstael 2005). They have been used for more than 30 years to evaluate a variety of economic policies (Hufbauer and
Elliott 1994, Winston 1993). Applications to the valuation of ecosystem services include Barbier and Strand (1998) and Barbier, Strand, and

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Sathirathai (2002). EPA has used these methods to value ecosystem service benefits from air pollution control in the markets for agricultural

products and for timber products (US EPA 1999).

Limitations

Estimating both consumer and producer surplus requires the development of empirical models for the demand and supply
relationships describing market outcomes. Depending on each application this can be difficult due to lack of data at the level of resolution
required to describe how economic policies affect each of these relationships.

The majority of environmental policies do not directly impact the prices and quantities of goods and services traded in markets, so this
method is only available in a limited subset of cases. In addition, it will only capture the benefits of a change that are manifested in marketed
outputs. For example, a wetland may contribute not only to commercial fishery production but also to flood control, water purification,
wildlife habitat, etc. These other benefits would not be captured by a market-based approach. Another limitation of this method is that, if
there are market imperfections stemming for example from market power, this can confound the measurement of demand and supply and
distort the relationship between prices and the marginal value and marginal cost of providing a private good. As a result, this distortion will
carry over into any estimation of economic values based on market prices.

Many non-environmental factors can affect demand and supply relationships that are also important. Seasonal variations in use or
availability of goods and services related to environmental policies can affect prices, and this needs to be considered. The modeling and
estimation of demand and supply functions can be complicated. Ultimately, what can be learned about the influence of environmental or any
other policy is limited by the available data. These limitations are best described as an identification problem - do we have sufficient
information to identify the effects that are hypothesized to reflect how environmental policy influences market supply and demand?

Key References

Barbier, Edward B., and Ivar Strand. 1998. "Valuing Mangrove-Fishery Linkages," Environmental and Resource Economics, 12:151-166.

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1	Barbier, Edward B., Ivar Strand, and Suthawan Sathirathai. 2002. "Do Open Access Conditions Affect the Valuation of an Externality?

2	Estimating the Welfare Effects of Mangrove-Fishery Linkages in Thailand, Environmental and Resource Economics, 21:343-367.

3	Boardman, Anthony E., David H. Greenberg, Aidan R. Vining, and David L. Weimer. 2006. Cost-Benefit Analysis: Concepts and Practice,

4	third edition Upper Saddle River, NJ: Prentice-Hall.

5	Hufbauer, Gary, and Kimberly Ann Elliott. 1994. Measuring the Costs of Protection in the US, Washington, DC: Institute for International

6	Economics.

7	McConnell, Kenneth E., and Nancy E. Bockstael, 2005. "Welfare Theory and Valuation," in Karl-Goran Maler and Jeffrey R. Vincent, eds.,

8	Handbook of Environmental Economics, Amsterdam: Elsevier.

9	US EPA. 1999. The Benefits and Costs of the Clean Air Act 1999 to 2010, Washington, DC.

10	Winston, Clifford. 1993. "Economic Deregulation: Days of Reckoning for Microeconomists," Journal of Economic Literature.

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Non-market Methods - Revealed Preference

When environmental changes affect goods and services that are not traded in markets, non-market valuation, using either revealed
preference or stated preference, becomes necessary. Revealed preference methods look at people's behavior in markets that are related to
ecological services to reveal underlying values. For example, someone's decision about which of two houses to purchase might reveal
information about how they value air quality or a scenic view if the two houses vary with regard to that environmentally-related attribute.
Because the revealed preference methods for measuring values use data on observed behavior, some theoretical framework must be
developed to model this behavior and to relate the behavior to the desired monetary measures of value and welfare change. A key element in
the theoretical framework is the model of the optimizing behavior of an economic agent (individual or firm) that relates the agent's choices to
the relevant prices and constraints, including the level of ecological services being provided. If a behavioral relationship between observable
choice variables and the ecosystem service can be specified and estimated, this relationship can be used to calculate the economic value of
changes in these service flows. For example, one well-established behavioral relationship is that between the costs to individuals of visiting a
recreation site and the numbers of visits make to the site. See the discussion of the travel cost method that follows. If the numbers of visits
also varies systematically with the level of an ecosystem service provided by the site, then the value of the ecological service can be inferred
from these relationships.

The degree to which inferences about the value of a change in ecosystem services can be drawn from market observations, and the
appropriate techniques to be used in drawing these inferences, both depend on the way in which the ecosystem service enters individual utility
functions. The exploitation of possible relationships between environmental goods and private goods leads to several empirical techniques
for estimating environmental and resource values. This section covers three revealed preference methods: travel cost, hedonics, and averting
or mitigating behavior models.

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Travel cost

Brief description of the method

The travel cost method accepts as a maintained hypothesis that people have economic demand functions for the services of
environmental resources that are associated with observable choices they make to travel to a particular location. While in principle this
method could be applied to travel for a variety of purposes, in practice it is applied in the context of travel associated with outdoor recreation.
Lakes, rivers, forests, and beaches are examples of the types of resources involved. The essence of the method is recognition that users pay
an implicit price by giving up time and money to take trips to these areas for recreation. This recognition is important because most of the
public facilities for recreation in the United States do not have market determined fees for that use. The cost of a visit to a site is the out-of-
pocket costs of travel including any site admission fees, opportunity cost of travel time, and the opportunity cost of time on site.45

The values of ecosystem services are captured by the method to the extent they can be represented as factors that influence a person's
decision about where or how often to travel. For example, a measure of the availability of fish in a lake used for fishing would presumably
influence (along with other factors) a person's decisions about whether or how often to visit the site for fishing.

Until about the middle 1990's, the travel cost literature estimated travel costs for the simple case of a new site or loss of site. The loss
of an area (due to activities that eliminate its recreational value) is represented as "equivalent to" a price or travel cost change that is large
enough to cause all existing users to no longer take trips to the site. To use the travel cost method for more sophisticated environmental policy
choices, i.e., those that change the quality of recreational opportunities, analysts need to know how those quality attributes influence the
demand function for recreation. In practice, most economic models for recreation now use random utility models (RUM), which describe the
decision process associated with each individual selecting which recreation site among a number of alternatives to visit. A RUM framework
describes these choices as the result of a constrained optimization process: selecting the site that yields the maximum level of utility (or well-
being) that is possible given a person's constraints. The result can be expressed as a function of travel costs, site characteristics such as the
level of ecosystem services and the facilities to support specific activities (e.g., boat ramps, ski lifts, etc.), and users' attributes.

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Status as method

The travel cost methodology is based on well-established economic principles. There has been extensive use of this method in peer-
reviewed literature, dating to 1947 when Harold Hotelling first proposed it. There is less experience with using the method to estimate trade-
offs for a wide range of attributes of recreation sites. Assumptions are understood and documented. Meta analyses - Smith and Kaoru
(1990), Walsh, Johnson and McKean (1992), Rosenberger and Loomis (2000), Johnston, et al. (2003) and Johnston, et al. (2005) have
documented the performance of the model in different circumstances.

Measures of the economic value have been used in EPA's RIA analyses for regulations affecting recreation resources. A recent
example is the Phase III component of the 316B rule. The rule seeks to reduce impingement and entrainment of fish and other organisms
through power facilities' uptake of cooling water.

Strengths and Limitations

The primary data requirements of the travel cost methodology are as follows: data on people's usage of recreation sites; measures of
individuals' values of time and time constraints; information that allows measures of the environmental attributes of the resources used for
recreation to be linked to those resources; and information that describes the relationship between technical indexes of the attributes of
recreation sites and measures that users can be expected to understand and know.

The analysis requires technical training in micro-economic modeling of demand and extensive experience with micro-econometrics to
estimate recreation demand models. Less experience is required to use existing models to estimate economic values for changes in factors
hypothesized to affect people's recreation behavior.

Uncertainties

One important source of uncertainty in the travel cost model is the value of recreationists' time as a component of the cost of a
recreation trip. Randall has argued that for several reasons "travel cost is inherently unobservable" (1994, p. 88). The role of time in
explaining recreation demand and in valuing recreation visits and sites raises some thorny issues for both the standard travel cost and RUM
approaches of analysis. Clearly, time is an important variable in the analysis of recreation demand and value. However, numerical estimates

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of demand and value require either that the numerical value of the shadow price of time be known or that it be estimated from a model of the
choices made regarding the uses of time. A variety of models of choice and time are available in the literature. However, as yet, different
model structures yield quite different estimates of the shadow price of time, and there is no clear basis for preferring one model and its value
over other models. Until these issues can be resolved, estimates of recreation values should be presented as conditional upon a specific value
of the shadow price of time or a specific modeling approach regarding the role of time, and the uncertainty in the estimates that this implies
should be acknowledged. For more on this issue, see Freeman (2003, Ch. 13).

Key References

P.A. Champ, K.J. Boyle and T.C. Brown, editors, A Primer on Non-Market Valuation (Dordrecht: Klumer Academic 2003).

A.M. Freeman III, The Measurement of Environmental and Resource Values, second edition (Washington, D.C. Resources for the Future
2003).

Haab, T.C. and K.E. McConnell, 2002, Valuing Environmental and Natural Resources, Cheltenham, UK: Edward Elgar.

Johnston, Robert J., Elena Y. Besedin, and Ryan F. Wardwell, 2003, "Modeling Relationships Between Use and Nonuse Values for Surface

Water Quality: A Meta-Analysis," Water Resources Research, 39(12).

Johnston, Robert J., Matthew H. Ranson, Elena Y. Besedin, and Erik C. Helm, 2005, "What Determines Willingness to Pay per Fish? A

Meta-Analysis of Recreational Fishing Values," under review at Marine Resource Economics.

D.J. Phaneuf and V.K. Smith. 2005. "Recreation Demand Models," in K. Maler and J. Vincent, editors, Handbook of Environmental

EconomicsVol.il. Amsterdam: North Holland.

Randall, Alan. 1994. A Difficulty with the Travel Cost Method. Land Economics 70(l):88-96.

Rosenberger, R.S. and J.B. Loomis, 2000, "Using Meta-Analysis of Economic Studies: An Investigation of Its Effects in the Recreation
Valuation Literature," Journal of Agricultural and Applied Economics 32(3): 459-470.

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Smith, V. Kerry and Yoshiaki Kaoru, 1990, "Signals or Noise? Explaining the Variation in Recreational Benefit Estimates," American

Journal of Agricultural Economics 72: 419-433.

Walsh, R.G., D.M. Johnson, and J.R. McKean, 1992, "Benefit Transfer of Outdoor Recreation Demand Studies, 1968-1988," Water
Resources Research 28(3): 707-713

Hedonics

Brief description of the method

Hedonic methods seek to exploit possible relationships between demands for private goods and their associated bundle of
characteristics, including environmental characteristics. For example, the demand for a house depends not only on its physical attributes (e.g.,
total size, the number of bedrooms, etc.) but also on the surrounding environmental characteristics (e.g., air quality, proximity to beach, etc.)
When people select from among the set of available goods (e.g., available houses), the hedonic model assumes that they will choose the one
that is their most preferred given its price and attributes. In equilibrium, the set of prices for these differentiated goods will be structured so
there is no incentive for anyone to change their choices. The hedonic price function relating prices to characteristics is a reduced form
description of this equilibrium condition. The primary applications of this logic in the field of environmental economics involve housing
prices and the wage rates for jobs

Assuming that the price of a house reflects the attributes of that house, its property, neighborhood, and facilities that are "near" it, then
the hedonic price function reflects a buyer's marginal willingness to pay (WTP) for small changes in one of these attributes. This measure is a
single point estimate of the marginal value. The method does not provide the basis for measuring, without additional assumptions, any
economic benefits that are associated with a large change in one or more of these attributes. These attributes can include the structural
features of the house, its lot, and the characteristics that are conveyed to those living in the home because of its location. For example, if a
house is on the coast, residents can experience the coastal views, any beach related amenities, as well as any greater risk of damage that might

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arise from coastal hazards. If that feature is some aspect of an ecological service available to an individual because she lives in the house, the
model allows that incremental value of a change in that service to be estimated.

If the attribute measures a characteristic that can be related to a policy, e.g., proximity to a Superfund site before and after clean up,
then it is possible to describe a buyer's willingness to make trade-offs for small changes in that attribute. There are important qualifications
that must be considered in evaluating the results from these models. For example, to the extent the prices for homes near wetlands or in flood
zones are found to be related to (i.e., have a statistically significant association with) the measures that are used to isolate these features, then
there is indirect evidence that these features are recognized by buyers and sellers. This result follows because they contribute to the observed
equilibrium prices for the homes represented by the hedonic function. Relating such a recognition to a measure of the incremental value for
the change in services requires assumptions describing how changes in the variable that can be measured and included in the price function
relate to changes in the service of interest.

Extensive data are needed to estimate a statistical function that relates housing prices to housing characteristics that include
environmental attributes so that small changes in the quality or quantity of that environmental attribute can be related to small changes in
housing prices.

Status as a Method

The hedonic method has been widely used to evaluate site-specific amenities and disamenities. Examples of applications involve: air
pollution, noise pollution, proximity to water bodies, wetlands, coastal areas, and location of homes in hazardous areas such as earthquake or
flood zones. See Palmquist (2005) for a general overview of the literature and Smith and Huang (1995) for a meta-analysis of the studies of
air pollution and property values. This and other meta-analyses indicate clear support for the methods for applications where we can expect
buyers and sellers to have knowledge of the amenities.

Applications involving site attributes that might be more closely aligned with services of ecosystems are much more limited. Several
studies have investigated the effects of proximity to wetlands of different types as well as for distance to open space. Examples include

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Mahan, et al. (2000), Netusil (2005), and Smith, et al. (2002). An important difficulty in using these results arises in converting the
incremental value estimated for a change in distance to a measure more directly related to changes in ecosystem service.

Strengths and Limitations:

Hedonic methods are familiar to most people who have purchased or sold a house because realtors do an informal hedonic type
analysis comparing homes described as "comparables" to price a proposed new listing.

The main strength of the hedonic housing method is that it is based on people's actual choices. However, all hedonic methods face
significant econometric hurdles and are subject to the standard criticism of statistical relationships that they reveal correlation but fall short of
revealing causation. Hedonic estimates can be sensitive to the choice of model specification (see, for example, Cropper, Deck and
McConnell 1988). Moreover, relating housing prices to many ecosystem services remains elusive. Finally, hedonic methods can only
capture the value of environmental changes that individual homeowners recognize. The method is best suited for local housing markets.
While several studies have estimated national hedonic property value models, it is generally agreed that it is unreasonable to assume that there
is a single national market for housing with an equilibrium that adequately describes the trade-offs among housing attributes in very different
locations.

To implement the method for estimating the hedonic price function, it is important to have access to a real estate transaction database
with sales prices, housing characteristics, and the latitude/longitude coordinates for each property. These data can then be merged to GIS files
describing access to various spatially delineated environmental resources such as air quality as well as to ecosystem services.

Uncertainty

The primary sources of uncertainty with the hedonic model for policy applications arise with the measurement of attributes that are
assumed to represent the environmental services available to people due to living in the house. Further research on how people learn about
these aspects of a location and what they consider to be conveyed by a location would help to address this issue.

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In addition, simulation analysis evaluating the performance of hedonic price functions as approximations to an equilibrium matching

process would also contribute to our understanding of the sensitivity of the method to assumptions about model structure and functional form.

See, for example, Cropper, Deck and McConnell (1988).

Key References

Champ, P.A., K.J. Boyle and T.C. Brown, editors. 2003. A Primer on Non-Market Valuation. Dordrecht: Kluwer Academic Press.

Cropper, M.L, L. Deck, and K.E. McConnell, 1988, "On the Choice of Functional Forms for Hedonic Price Functions," Review of
Economics and Statistics, 70: 668-75.

Freeman, A.M. III. 2003. The Measurement of Environmental and Resource Values, second edition. Washington, D.C.: Resources for the
Future).

Haab, T.C. and K.E. McConnell. 2002. Valuing Environmental and Natural Resources, Cheltenham, UK: Edward Elgar.

Mahan, B.L., S. Polasky, and R.M. Adams, 2000, "Valuing Urban Wetlands: A Property Price Approach," Land Economics, 76 (February):
100-113.

Netusil, Noelwah, 2005, "The Effect of Environmental Zoning and Amenities on Property Values: Portland Oregon," Land Economics, 81
(May): 227.

Palmquist, Raymond B., 2005, "Hedonic Models" in K. Maler and J. Vincent, editors, Handbook of Environmental Economics Vol. II
Amsterdam: North Holland.

.V. Kerry Smith, Poulos, Christine, and Hyun Kim, 2002, "Treating Open Space as an Urban Amenity," Resource and Energy Economics 24:
107-129.

Averting behavior models

Brief Description of the Method

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Averting or mitigating behavior models simulate consumer behavior and rely on the existence of an activity that substitutes for the
services provided by an environmental resource. The averting behavior method infers values from defensive, mitigating, or averting
expenditures, i.e., those actions taken to prevent or counteract the adverse effects of environmental degradation. For example, an individual
might purchase a water filter to avoid the health risks associated with drinking unfiltered water. By analyzing the expenditures associated
with these defensive purchases, researchers impute a value that individuals place on small changes in environmental or health risks. In effect,
a defensive expenditure is spending on a good that is a substitute for health protection or an environmental quality or service. Because the
method is based on an estimation of the marginal rate of technical substitution between the environmental service and a market good or
service with a known market price, it is capable of producing monetary estimates of the value of the environmental service. What is required
is an understanding of the technical relationships underlying the ability of the environmental service and its market good substitute to enhance
human well-being.

Status of the Method

There is a substantial literature on the theoretical dimensions of the method (for example, Freeman 2003, Dickie 2003, Smith 1991)
but relatively few convincing studies demonstrating it will work in practice. Examples of defensive expenditures include the choice of
automobile type (as it relates to fatality risk), safety helmets, fire alarms, and water filters. However, since these expenditures only capture a
portion of an individual's willingness to pay (WTP) for these protections, averting behavior results are sometimes interpreted as a lower
bound on willingness to pay to avoid a particular harm. The most common application of averting behavior models has been the estimation
of values for morbidity (illness) risk.

Limitations

Averting behavior studies rarely provide economic values for ecosystem services. Even for those averting behavior studies for water
quality, the motivation for the averting behavior is usually to protect health or life.

Key References

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1	Dickie, Mark. 2003. "Defensive Behavior and Damage Cost Methods," in Champ, P.A., K.J. Boyle and T.C. Brown, editors, A Primer on

2	Non-Market Valuation. Dordrecht: Kluwer Academic Press.

3	Freeman, A.M. Ill 2003 The Measurement of Environmental and Resource Values, second edition.Washington, D.C.: Resources for the

4	Future.

5	Smith, V. K. (1991), "Household Production Functions and Environmental Benefit Estimation," in J. B. Braden and C. D. Kolstad, eds.,

6	Measuring the Demand for Environmental Quality. Amsterdam: North Holland

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Non-market Methods - Stated Preference
Brief Description of the Method

Stated preference methods rely on survey questions that ask individuals to make a choice, describe a behavior, or state directly what
they would be willing to pay for specified changes in environmental services not traded in markets. The various stated preference techniques
are distinguished by how the information is presented, what questions are asked, and how their responses are formatted. It is important to
acknowledge that the choices, stated values, or revised patterns of use are derived from answers to questions that ask respondents what they
would do, or how much they would pay for, or how they would alter their choices in response to changes in the amount of a non-market good
or service in a specified hypothetical setting. This is in contrast to Revealed Preference Methods, which are based on observing the actual
choices made by people facing real constraints on income, etc. Stated preference methods offer the opportunity to measure trade-offs for
anything that can be presented as a credible and consequential choice. Hence, their primary advantage is their ability to, in principle, measure
a wider set of values. In particular, they are the only economic methods that can measure non-use values.

Although not all authors use the same terminology, the term stated preference methods generally include any survey questions in
which respondents are asked hypothetical questions designed to reveal information about their preferences or values. The term encompasses
three broad types of questions. The first type involves questions that ask directly about monetary values for a specified commodity or
environmental change. These are usually called contingent valuation method questions (CVM). In the past, the most commonly used CVM
questions simply asked people what value they place on a specified change in an environmental amenity or the maximum amount they would
be willing to pay to have it occur. These are usually open-ended in that the individual has to state a number rather than respond to a number
offered by the researcher. The responses to these questions, if truthful, are direct expressions of value. The other major type of CVM question
asks for a yes or no answer to the question, "Would you be willing to pay $X ...?" Each individual's response reveals only an upper bound
(for a no) or a lower bound (for a yes) on the relevant welfare measure. Questions of this sort are termed discrete choice questions.

Responses to discrete choice questions can be used to estimate willingness-to-pay functions or indirect utility functions.

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The second and third major types of Stated Preference methods do not reveal monetary measures directly. Rather, they require some
form of analytical model to derive welfare measures from responses to questions. The second type of question is called variously "choice
experiment," "conjoint analysis," or sometimes an "attributes based method" (Holmes and Adamowicz 2003). In this approach to
questioning, respondents are given a set of hypothetical alternatives, each depicting a different bundle of environmental attributes.
Respondents are asked to choose the most preferred alternative, to rank the alternatives in order of preference, or to rate them on some scale.
Responses to these questions can then be analyzed to determine, in effect, the marginal rates of substitution between any pair of attributes that
differentiate the alternatives. If one of the other characteristics has a monetary price, then it is possible to compute the respondent's
willingness to pay for the attribute on the basis of the responses.

In the third type of SP question, individuals are asked how they would change the level of some activity in response to a change in an
environmental amenity. If the activity can be interpreted in the context of some behavioral model such as an averting behavior model or a
recreation travel cost demand model, the appropriate indirect valuation method can be used to obtain a measure of willingness to pay. These
are known as contingent behavior or sometimes contingent activity questions.

Status of the Method

The method has an extensive literature of principles and applications extending over a forty-year period. Mitchell and Carson's
(1989) pioneering treatise is still the primary reference on CVM, especially for design and implementation questions. See also Carson (1991).
Two new works that focus on best practice and empirical estimation for CVM and stated choice studies are Boyle (2003) and Holmes and
Adamowicz (2003), respectively. The so-called NOAA Blue Ribbon Panel (U.S. National Oceanic and Atmospheric Administration 1993)
reviewed CVM in the context of assessing damages to natural resources in support of litigation and provided its guidelines for best practice.
Other important references are: Bjornstad and Kahn (1996) for a review of theoretical and empirical issues that includes assessments by both
proponents and critics of stated preference methods; Kopp, et al. (1997); Bateman and Willis (1999); Bateman, et al. (2002); and Smith
(2004, 2007).

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Use of the stated preference methods for environmental valuation has been controversial. A major issue concerning the status of
stated preference methods is the validity of the resulting value estimates. There are several concepts of validity and various approaches to
assessing the validity of responses. A commonly cited issue related to validity is the existence of what is known as hypothetical bias. The
argument is that the hypothetical nature of stated preference questions results in the overstatement of economic values, or what is known as
hypothetical bias. However, the evidence regarding the extent of this bias is mixed (see Murphy, et al. 2005 for a recent discussion). The
controversy surrounding stated preference methods had the salutary effect of stimulating a substantial body of new research on both practice
and on the credibility or validity of stated preference estimates of value. A good overview of the issues raised in this controversy is contained
in the three essays published as a symposium in the Journal of Economic Perspectives (Portney 1994, Hanemann 1994, and Diamond and
Hausman 1994). See also, Hausman (1993) and Freeman (2003) and references therein for further discussion.

Strengths and Limitations

Strengths include the accumulated experience of forty years of practice and research. Also in principle, stated preference methods are
the only set of methods capable of capturing so-called nonuse values, since without use there is no behavior that can reveal values through
application of revealed preference methods.

In addition to the controversy stemming from the hypothetical nature of the questions noted above, some people question whether
surveys are capable of providing useful information about preferences. One issue is whether preferences regarding unfamiliar environmental
goods are well-formed and stable (see Part 1, section 2.4). In addition, since responses to questions must reflect in some sense the knowledge
that individuals have about the thing being valued as well as respondents' preferences, the methods cannot be used to value ecosystem
services about which people are ignorant. For example, if respondents were asked questions concerning phytoplankton but were ignorant of
the role of phytoplankton in supporting the aquatic food chain and higher order species that they might value, their responses might be
interpreted as placing no value on phytoplankton. In such a case, stated preference methods will not generally be useful for valuing changes
in supporting ecosystem services (see Part 1, section 2.1) since most lay individuals are not aware of the crucial role of these services. One
solution to this problem is to use the survey instrument to convey information to respondents about the role of the ecosystem service being

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valued and the potential consequences of changes in the level of this service. See, for example, Banzhaf, et al. (2004). Then, of course, the
question becomes one of the validity of the information provided to respondents and the potential for biasing responses by providing biased
information.

Finally, even if preferences are well-formed and individuals are aware of the role of the relevant environmental attributes, the survey
might not provide incentives for respondents to reveal their preferences accurately. This depends, among other things, on the degree of
incentive compatibility of the various questioning formats and the set of methods as a whole. Carson, et al. (2000), reasoning from first
principles about what is in the best interest of respondents faced with a scenario, payment vehicle, and elicitation question, have established
under what conditions stated preference questions give people incentives to reveal their true values. The first two conditions are that the
survey question be about something that matters to the respondent and that the respondent believes that his/her response might affect the
outcome of the policy issue that is the subject of the survey. If both conditions hold, then the survey question is termed "consequential" to
respondents. For consequential questions, it is possible to reason from an assumption of acting on rational self interest to predict whether
responses will be truthful and if not, then at least in some cases what the direction of bias will be.

For consequential questions, the only question format that can in principle be incentive compatible is the single discrete choice
question. In addition, this form requires the further condition that the government agency is perceived as being able to compel payment of
some amount from the respondent if the good is provided. For example, questions that ask about the willingness to make a voluntary
contribution to support some government action fail this condition and provide incentives to respond "yes" even when the requested
contribution is greater than the respondent's WTP.

Key References

Banzhaf, Spencer, Dallas Burtraw, David Evans, and Alan Krupnick. 2004. Valuation of Natural Resource Improvements in the

Adirondacks. Washington, DC: Resources for the Future.

Bateman, Ian J., et al. 2002. Economic Valuation with Stated Preferences: A Manual. Cheltanham, UK: Edward Elgar.

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Bateman, Ian J., and Kenneth G. Willis (eds.). 1999. Valuing Environmental Preferences: Theory and Practice of the Contingent Valuation

Method in the US, EU, and Developing Countries. Oxford, UK: Oxford University Press.

Bjornstad, David J., and James R. Kahn (eds.). 1996. The Contingent Valuation of Environmental Resources: Methodological Issues and

Research Needs Cheltenham, UK: Edward Edgar.

Boyle, Kevin J. 2003. Contingent Valuation in Practice. In A Primer on Non-market Valuation edited by Kevin J. Boyle and Patricia A.

Champ. Boston: Kluwer Academic Publishers.

Carson, Richard T. 1991. Constructed Markets. In Measuring the Demand for Environmental Quality, edited by John Braden and Charles

Kolstad. Amsterdam, The Netherlands: Elsevier.

Carson, Richard T., Theodore Groves, and Mark J. Machina. 2000. Incentive and Informational Properties of Preference Questions.

Unpublished. http;//weber.ucsd.edu/~rcarson/ (accessed on August 20, 2002).

Diamond, Peter A., and Jerry A. Hausman. 1994. Contingent Valuation: Is Some Number Better Than No Number? Journal of Economic
Perspectives 8(4):45-64

Freeman, A. M. III. 2003. The Measurement of Environmental and Resource Values: Theory and Methods. 2nd ed. Washington, DC:
Resources for the Future

Hanemann, W. Michael. 1996. Valuing the Environment Through Contingent Valuation. Journal of Economic Perspectives 8(4): 191-43.
Hausman, Jerry A., (ed.). 1993. Contingent Valuation: A Critical Assessment. Amsterdam: North-Holland.

Holmes, Thomas, and Wictor Adamowicz. 2003. Attribute-Based Methods. In A Primer on Non-market Valuation edited by Kevin J. Boyle

and Patricia A. Champ. Boston: Kluwer Academic Publishers.

Kanninen, Barbara J. editor 2007 Valuing Environmental Amenities Using Stated Choice Studies (Dordrecht, The Netherlands: Springer)
Kopp, Raymond J., Werner W. Pommerehne, and Norbert Schwarz (eds.). 1997. Determining the Value of Non-Marketed Goods: Economic,
Psychological, and Policy Relevant Aspects of Contingent Valuation Methods. Boston: Kluwer Academic Publishers.

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Mitchell, Robert Cameron, and Richard T. Carson. 1989. Using Surveys to Value Public Goods: The Contingent Valuation Method.

Washington, D.C.: Resources for the Future.

Murphy, James J., P. Geoffrey Allen, Thomas H. Stevens, and Darryl Weatherhead. 2005. "A Meta-Analysis of Hypothetical Bias in Stated

Preference Valuation," Environmental and Resource Economics, 30(3):313-325.

Portney, Paul R. 1994. The Contingent Valuation Debate: Why Economists Should Care. Journal of Economic Perspectives 8(4):3-17.

Smith, V. Kerry 2004, "Fifty years of Contingent Valuation" in T. Tietenberg and H. Folmer editors, International Yearbook of

Environmental and Resource Economics 2004/2005 (Cheltenham, U.K. Edward Elgar) ppl-60.

Smith, V. Kerry 2007 "Judging Quality" in B. Kanninen editor Valuing Environmental Amenities Using Stated Choice Studies (Dordrecht,
The Netherlands: Springer)

U. S. National Oceanographic and Atmospheric Administration. 1993. Report of the NOAA Panel on Contingent Valuation. http://web.lexis-
nexis.com/congcomp/printdoc (accessed on August 12, 2002).

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Combining Revealed and Stated Preference Methods

It is possible to combine revealed and stated preference methods to estimate what both types of choices imply for characterizing an
individual's willingness to pay for changes in environmental services. Cameron (1992) was the first to propose this idea for environmental
applications. To be informative, this strategy must be based on an analysis of the revealed and stated behaviors to establish that the empirical
models used to describe these outcomes share at least one parameter. That is, they must each be capable of identifying at least one common
parameter. Ideally there would be more parameters shared between the models. Most applications collect the two types of data (i.e., revealed
and stated preference) from the same respondents. This requirement is not essential. It would be possible in principle to combine samples
with different respondents providing the revealed and stated components of the analysis. A key issue in applying these methods to the task of
valuing ecosystem services is the need to have measures for the quality and amount of ecosystem services that are compatible with models
and data typically available for revealed and stated preference models.

See Adamowicz, et al. (1994), Earnhart (2001, 2002), and McConnell, et al. (1999) for more recent applications.

Key References

Adamowicz, W., J. Louviere, and M. Williams. 1994. Combining Revealed and Stated Preference Methods for Valuing Environmental

Amenities. Journal of Environmental Economics and Management 26(3):271-292.

Cameron, Trudy A. 1992. Combining Contingent Valuation and Travel Cost Data for the Valuation of Nonmarket Goods. Land Economics
68(3):302-317.

Earnhart, Dietrich. 2001. Combining Revealed and Stated Preference Methods to Value Environmental Amenities at Residential Locations.
Land Economics 77(1): 12-29.

Earnhart, Dietrich. 2002. Combining Revealed and Stated Data to Examine Housing Decisions Using Discrete Choice Analysis. Journal of
Urban Economics 51(1): 143-169.

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McConnell, Kenneth E., Quinn Weninger, and Ivar E. Strand. 1999. Joint Estimation of Contingent Valuation and Truncated Recreational
Demands. In Valuing Recreation and the Environment: Revealed Preference Methods in Theory and Practice, edited by Joseph A.
Herriges and Catherine L. Kling. Cheltanham, UK: Edward Elgar.

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GROUP EXPRESSION OF VALUES AND SOCIAL/CIVIC VALUATION

Valuation of ecological systems can also involve expressions of group or public value, rather than elicitations of the values of
individuals or biophysical rankings according to a previously agreed-upon scale. Group or public expressions of ecological value have
attracted attention for at least two reasons. First, some experts believe that group discussions and deliberations can help people form clearer
understanding of values. Second, a number of experts believe that group expressions of the "public good" in general, and of ecological value
in particular, may be distinct from the aggregation of individuals' reports of their private welfare because they explicitly reflect public
regardedness.

Although many reports briefly discuss the potential role of deliberative processes in helping to develop more informed valuation
(National Research Council 2004, Millennium Ecosystem Assessment Board 2003, Science Advisory Board 2000), the reports do not
evaluate or recommend any specific method or approach. The committee notes parallels between group and public expressions of value for
ecological valuation and the deliberative-analytic process recommended for risk characterization by the National Research Council (1996).
The National Research Council report, however, did not address in any detail how deliberative approaches might be implemented or assessed
or how they might be transferred to ecological valuation.

Traditional economic valuation methods attempt to measure and aggregate the values that individuals place on changes in ecological
systems and services based on their personal preferences as consumers of those systems and services. An alternative approach is to try to
measure the values that groups of individuals place on changes in such systems and services explicitly in their role as citizens - social/civic
valuation. This approach measures the monetary value that groups place on changes in the systems and services when asked to evaluate how
much the public as a whole should pay for increases in such systems and services (public willingness to pay) or should accept in
compensation for reductions in the systems and services (public willingness to accept). The value measurement purposefully seeks to assess
the full "public regardedness" value, if any, that the group attaches to any increase in community well-being attributable to changes in the
relevant systems and services.

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Social/civic values, like values based on personal preferences, can be measured either through revealed behavior or through stated
valuations. One principal source of revealed values for changes in ecological systems and services are votes on public referenda and
initiatives involving environmental decisions. Other public decisions also may provide measures of social/civic values, including official
community decisions to accept compensation for permitting environmental damage and jury awards in cases involving damage to natural
resources. Because all research on sources of revealed public value have focused on referenda and initiatives, however, this section discusses
only the use of referenda and initiatives as a source of revealed value. Other public decisions raise unique issues as sources of revealed value.
The committee does not recommend that EPA currently pursue their development. Where revealed values are difficult or impossible to
obtain from referenda or initiatives, social/civic values may be measured by asking "citizen valuation juries" or other representative groups
the value that they, as citizens, place on changes in particular ecological systems or services.

This section discusses several approaches to forming, eliciting and considering group or public values. Some of the methods are
designed to help elicit clearer understandings of value, while others focus on identifying group expressions of public valuation. The
committee recommends each method be considered for its merits at different stages in the ecological valuation process and in difference
decision-contexts relevant to EPA

Method Form iil'iiMipui uiuls'.' \Vh;ii is method intended In incisure'.'

Source of Information About Value







Does method measure obsened heha\ lor.
\ erhal or u nlieu expressions, or progress
related lo pre\ iousl\ identified goal'.'

Who

expresses
Milne-.'

Focus Groups

Narrative summaries, frequency
tallies, consensus

Full discovery and articulation of all
the values that are relevant and
exploration of agreements and
conflicts among stakeholder
constituencies

verbal reports

sample from
public

Referenda and
Initiatives

Historical monetary data on
communities' choices regarding
ecological impacts

What the body politic as a collectivity
values in terms of policy outcomes

Behavior

Selected
stakeholders

Citizen Valuation

Qualitative summary of jury

How a representative group views the

Verbal reports

Selected

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\ lei hud

I'OI'lll ol'oiiiput lllllls'.'

\Vlial is method intended In measure'.'

Source of liilornKilion About Value







Does method measure obsened lvha\ lor.

Who







\ erhal or u mien expressions. or progress

expresses







relaled lo pre\ loush identified izoal'.'

Millie"

Juries

decisions which may include
quantitative or monetary
decisions

social civil value of changes to
ecological systems and services



stakeholders

1



Degree lo u Inch Method 1 las
1 icon l)e\ eloped or I tili/ed

Recommendations lor Research to Sircuuthen I se of

Potential lor I'uliirc I se b\ HP A
m an Integrated and 1 \panded
Approach lor \ alualioii

Issues lu\ ol\ cd iu
Implementation

Focus Groups

•	Not clear the extent to
which focus groups are
systematically used in
EPA policy making

•	The OMB and other
guidelines do not clearly
specify the criteria for
using focus groups



•	Can be useful and important
for designing and pre-testing
more formal surveys

•	May also contribute to the
design of more effective
communications of Agency
decisions



Referenda and
Initiatives

• Logic has been used
primarily in the literature
on health and safety

• The research needed to make the results of public
decisions through referenda and initiatives most
useful for inferring values would consist of the
creation of a data bank of referenda and initiative
outcomes, optimally screening out those involving
multiple, confounding elements.

•	Can provide monetized
values—of the community's
formal decision and values,
ceilings, or floors of the
median voter's valuation

•	With follow-up surveys can
provide information on
beliefs, assumptions and
motives regarding the
ecosystem preservation
issues that voters perceive
are at stake

•	Any EPA decision context
calling for monetized
valuation could employ
these variants, either singly
or as cross-checks with
conventional revealed

• Analysis meets the
criteria for when
method "works best"

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Deurce id u Inch Method 1 las
1 icon l)e\ eloped or I lili/ed

Recommendations I'm' Research u> Streimiheii I so of

Potential for 1 "uliiiv I so b\ I P \
in an Integrated and 1 \panded
Approach lor Valuation

Issues lu\ ol\ ed in
Implementation







preference or stated
preference approaches.



Citizen

Valuation Juries

•	Experimental method in
the context of ecological
valuation

•	Used primarily to help
governments rank options
for achieving particular
goals. Only a few efforts
have been made to date to
use citizen juries to
generate monetary or
other estimates of the
social/civic value of
environmental changes.

•	Do citizen valuation juries arrive at different
valuations than individual respondents to CV
surveys? If so, how and why do the valuations
differ?

•	How stable are valuations provided by citizen
juries? How much variation exists among the
valuations produced by different citizen juries?

•	How do jury selection processes affect the
valuations of the jury? What methods exist to
overcome the inevitable bias arising from the small
size of citizen juries?

•	How should information be provided to citizen
valuation juries?

•	How do decision making rules (e.g., consensus
versus unanimity) affect valuations? What are
relevant considerations in choosing among the
different decision making rules?

•	Potentially useful both to
identity socially important
assessment endpoints and to
attach a value, monetary or
socio-psychological, to
changes in the assessment
endpoints

•	Can expand the role that the
public plays in valuations of
changes in ecological
systems and service

•	Hypothetical
character of all stated
preference valuations

•	Issues of group
dynamics

•	Choice of jurors

1

2

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Focus Groups
Brief description

Focus group methods engage small groups of relevant stakeholders in facilitated discussion and deliberation on selected/focused
topics relevant to the assessment of the effects of a policy, or alternative policies, outcomes, and/or consequences. Typically, experts and/or
trained facilitators present the context, motivation and goals for the group and open-ended narratives are collected from the participants,
usually in the context of discussion and deliberation with other members of the group and the experts/facilitators. Collected narratives are
subjected to qualitative analyses to identify and possibly to ascertain levels of consensus on relevant issues, perspectives, and positions
represented by the participants. Reports of focus group results typically include numerous quotations of collected comments, along with the
investigators' interpretations of the implications for the problems/policies/outcomes being addressed (e.g., Winter and Fried 2000). Less
often collected narratives are subjected to more rigorous analyses based on formal logic models or discourse analysis systems (Abell 2004,
Bennett and Elman 2006).

Relative to formal surveys, focus groups use small numbers of respondents and do not typically attempt formal probability sampling
to select participants. Emphasis is instead on assuring that at least one representative from the full range of interests and perspectives relevant
to the policies or outcomes at issue are included. The goal of a focus group is rarely value assessment per se, but a full discovery and
articulation of all of the values that are relevant, and exploration of agreements and conflicts among the stakeholder constituencies
represented by participants. Thus, focus groups are often employed early in policy and decision making, including the identification of the
problems to be addressed and the formulation of alternative policies to address those problems. It is common for focus groups to be used in
the process of designing and pre-testing more formal surveys. For example, Shields, et al. (2002) reported that 80 focus groups distributed
across the nation were used in developing the USDA Forest Service survey illustrated in Text Box 12.

Relation of Method to the C-VPESS Expanded and Integrated Assessment Framework

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Focus groups would be most appropriate and most useful at the earliest and latest stages of the decision-making process. While focus
groups do not generally provide quantitative assessments for alternative policies or outcomes, they can make important contributions to
improving the design, development, and pre-testing of more formal surveys that can provide reliable and valid quantitative assessments of
public concerns and values. Genuine probing interactions with individuals or groups representing key stakeholders and including divergent
views and concerns should be a central part of problem definition and identification of significant ecological and associated social effects
components of the process. Such interactions with key stakeholders and with citizens could also inform the values integration and negotiation
in the final decision process and guide and pre-test the communication of that decision.

Status of Method

It is not clear the extent to which focus groups are systematically used in EPA policy making, nor do the OMB and other guidelines
clearly specify the criteria for using these methods. Focus groups are widely used in marketing and political polling contexts and the U.S.
Forest Service national survey by Shields, et al. (2002) described above reported that "over 80 focus groups conducted around the continental
United States" (p. 1) were used in the design and development of the survey, as well as to support the interpretations and conclusions from the
survey. Public meetings and on-site demonstrations are frequently cited as playing a public involvement role in EPA policy decisions, and a
formal "Multi-Stakeholder Group" was assembled and used in the Avtex Fibers Superfund Site decision and implementation process, but it is
not clear whether any of these activities can be construed as using a focus group, nor is it clear how often such methods have been used to
systematically compare alternative policies/actions.

The use of focus groups would seem to be completely consistent with previous advice of the EPA Science Advisory Board (US EPA
2001) recommending increased use of "stakeholder processes" in Agency decision making. Stakeholder processes were defined as ".. .group
processes in which the participants include non-expert and semi-expert citizens, and/or representatives of environmental non-governmental
organizations, corporations and other private parties in which the group is asked to work together to: define or frame a problem; develop
feedback in order to better inform decision makers about proposed alternative courses of action; develop and elaborate a range of options
and/or criteria for good decision-making which a decision-maker might employ; or, either explicitly or implicitly, actually make

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1	environmental decisions." (p 8) Still, the term "focus group" was not used anywhere in this document. While no specific evidence has been

2	found either way, it seems reasonable to assume that individual narrative interviews have not been important components of EPA decision-

3	making processes. Certainly the qualitative nature of the information provided by both focus groups and individual interviews, and the

4	general disinterest in representative sampling makes them poor candidates for formal policy evaluation exercises, but that does not preclude

5	their having a role in earlier stages of the decision-making process as envisioned by the C-VPESS.

6	Focus groups can have important and useful roles to play in Agency policy and decision making. However, their emphasis on

7	qualitative analyses and their typical disregard for representative sampling can make them less useful for systematic evaluations or

8	comparisons of alternative policies and outcomes. The method can be very useful and important for designing and pre-testing more formal

9	surveys that do provide quantitative assessments of values for alternative policies and outcomes. Qualitative methods may also contribute to
10	the design of more effective communications and rationalizations of Agency decisions to stakeholders and to the general public.

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Referenda and Initiatives
Brief description of the method

Referendum and initiative votes provide the basis for a set of valuation approaches that can yield monetized values, but use somewhat
different logic than that of the conventional individually based revealed-preference and stated-preference methods. The outcomes of
referenda (measures placed on the ballot by a legislative body) and initiatives (ballot measures proposed by citizens) directly express what the
body politic collectivity values in terms of policy outcomes. These expressions may or may not correspond closely to the aggregated values
of the individuals in the community in terms of outcomes. Referenda approaches (not to be confused with the "referendum format" often used
for posing questions to solicit contingent valuation responses) provide information about the policy preferences of the median voter; under
certain circumstances this information can tell us about the median voter's valuation of specific environmental amenities, and can even
provide information, albeit weaker, about mean valuations of those who participate in the voting process. They can also be useful for cross-
validating any other valuation approach that permits a prediction as to the outcome of a referendum or initiative. When a referendum or
initiative is followed by a survey to determine what voters believed the financial burden to be, the approach can also elicit relevant beliefs and
motives to reinforce the specific willingness-to-pay or willingness-to-accept information.

There are four variants for analyzing referenda and initiatives:

•	Referendum/initiative analysis

•	Analysis of public decisions to accept pollution or resource depletion

•	Referendum/initiative analysis followed by a survey

•	Analysis of public decisions to accept pollution or resource depletion followed by a survey

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Direct referendum/initiative analysis, with or without a follow-up survey, can evaluate trade-offs between community and/or
household costs (higher taxes, possibly job losses) and eco-system improvements (establishment or improvement of air, water, biodiversity
protection, etc.). Direct analysis of public decisions to accept pollution or resource depletion, with or without a survey, can evaluate trade-
offs between community and/or household benefits (increase in tax base, job creation, infrastructure improvements, etc.) and eco-system
deterioration (greater pollution, amenity reductions).

Tc\l l$o\ 14: Diivcl An;il\sis of Public Decisions lo Accept Pol In I ion or Resource Depletion

Some public \ otes can pro\ ide inferences lor willingness-to-accept decisions. These decisions in\ ol\ e a community's \ ole as lo u helher lo
permit the entry of a new linn or a new (or increased) economic acti\ iiy despite the expectation that such permission will degrade the
ecosystem The payment represents the ceiling on the community's \ aliiation of the en\ironmental amenities that are being relinquished It is
a ceiling because of the possibility that the community would ha\e accepted a lower le\el of compensation, and if the community \ allied the
forgone eco-system ser\ ices more than the compensation, then presumably it would not ha\ e accepted the compensation I lowe\ or. if there is
a \ ole and the outcome is close, the calculated \alualion can be considered to be close to the community's \aluation

The estimation task in\ ok es assessing the amount of en\ ironmental damage in physical terms and llie amount of compensation in monetary
terms Typically this compensation will come in the form of additional sources of taxes, the \alue of infrastructure that the new entrants
pro\ ide lor the community, additional income earned by community members, elc The per-househokl as well as per-community
compensation would be rele\ant I or example, the entry of an air-polluting factory may be accepted only afler the factory's owner commits lo
a certain number of jobs for the community, building a park, upgrading roads, contributing to the community's \ocational program.

()b\ iously many "community decisions" to permit the entry of polluters or other acti\ ities that degrade the ecosystem are not amenable to this
approach, because community leaders negotiate the le\el of benefits that llie community will receke without a \ote being taken, or the
benefits or costs are difficult lo estimate

Tc\l l$o\ 15: Kcrcmi(liim/lnili;ili\c \n;il>sis hollowed In ;i Sur\c>

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The alternative to relying solely on the referendum or initiative outcomes to make willingness-to-pay estimates consists of combining the
voting outcome with a follow-up survey to determine the perceptions of the stakeholders. This variant amounts to a hybrid of the first variant
and the "referendum format" contingent valuation approach. The floor of the willingness-to-pay value of the proposed eco-system
improvements is estimated by determining the voters' perceptions of the eco-system improvements and costs proposed by a recent
referendum or initiative. The respondents are asked whether they voted, how they voted, and what they believed the benefits and costs of the
proposal were. The quantitative analysis of results of the referendum/initiative is the same as direct analysis without a survey, but using the
perceived rather than actual stakes.

If, in addition to asking how respondents voted and their perceptions of the benefits and costs of the proposal, the randomly-sampled
respondents who opposed the proposal are asked what (lower) cost would have induced them to vote for the proposal, and those who
supported the proposal are asked how much more they would have been willing to pay, this approach also permits an estimate of aggregate
and mean values, just as a standard contingent valuation study would, with less potential distortion arising from respondents' desire to be
regarded in a favorable light. Thus the survey following a referendum or initiative can provide an internal cross-check of how much
correspondence there is between the stated-preference approaches and the referendum or initiative findings (Schlapfer, Roschewitz, &
Hanley 2004, Vossler and Kerkvliet 2003). In fact, the voting results can serve as a cross-check for any of the survey or other individual or
group assessment methods.

It should be noted that in focusing on the benefits and costs that respondents report, rather than the actual benefits and costs that the
referendum or initiative proposal specifies, the results do not reflect the community's formal decision. This is a significant difference in the
philosophy underlying the standing of the results. That is, the first variant, even if it does not necessarily reflect the values that voters
perceive, does represent what the voters have chosen. On the other hand, without the survey, the analyst cannot be certain what financial
impact the voter believes is at stake, inasmuch as many initiatives and referenda do not explicitly specify the voter's financial burden.
Different logics underlie their standing.

Text Box 16: Public Decisions to Accept Pollution or Resource Depletion Followed by a Survey

Just as the analysis of referendum and initiative outcomes can be augmented by determining voters' perceptions of the stakes, the ceiling of
the willingness-to-accept value of eco-system deterioration can be estimated by determining the benefits perceived by voters who supported
the arrangement accepting the entry of a polluting or depleting operation into the community, and their perceptions of the damage that would
be done. Like the direct analysis of willingness-to-accept votes, if the arrangement was approved by the electorate, and the property rights are
clear and transactions are low, the ratio of the perceived costs and compensation represents the ceiling of the median voter's valuation. The
survey, best administered as soon as possible after the actual vote, would reveal what the community members interpreted the benefits and

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costs Id he. lluis bringing the \ aluation closer lo iiuli\ idual \ allies linl again. there exists a trade-off that the results would not ha\ e standing
as the "community"s choice " If the sur\e\ includes the questions of the comenlional contingent \aluation regarding how much each
respondent would ha\e heen willing to accept, then the results would he e\en more robust in finding mean and aggregate \aluations as well as
median \aluations

How the method could be used as part of the C-VPESS expanded and integrated framework

These public decision approaches can provide monetized values—of the community's formal decision and values, ceilings, or floors
of the median voter's valuation. In addition, with the follow-up surveys they can provide information on beliefs, assumptions, and motives
regarding the ecosystem preservation issues that the voters perceive are at stake. Because the approaches focus on the content of proposals
before the voting public, they do not directly identify ecosystem service impacts as a natural scientist or engineer would, but they will reflect
voters' assessments of ecosystem service impacts. The approaches focusing exclusively on the decision outcomes do not directly identify
changes in ecosystems and ecosystem services that are of greatest concern to people, although the survey variants can include questions to
elicit this information. The approaches do address ecological impacts that other monetized approaches may underestimate, in that
participation in citizenship, in contrast with the private-utility decisions reflected in the standard revealed-preferences approaches, can reflect
concern for community well-being ("public regardedness") insofar as voters hold such regard. The approaches do not involve inter-
disciplinary collaboration among physical/biological and social scientists or ecologists. There is a very strong potential that a data bank of
inferred values from fairly large numbers of referenda and initiatives would assist EPA in presenting ranges of value for benefit transfers.
Status as a method

The logic of using formal public outcomes to infer how much society values particular outcomes has been used primarily in the
literature on health and safety. For example, the value of a "statistical life" has been estimated by calculating how much public policies
commit to spend in order to reduce mortality rates from health or safety risks, or, conversely, how much economic gain is associated with
public decisions that reduce safety (e.g., by examining official decisions of U.S. states to raise or lower speed limits, Ashenfelter &
Greenstone [2004] estimated the market value of the time saved by getting to the destination more quickly, and from that estimated the value

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of the additional expected traffic fatalities). The logic of making valuation inferences from referenda and initiatives has been addressed in a
few publications, most directly in Deacon & Shapiro (1975) and Shabman & Stephenson (1996).

In comparing the valuations yielded by stated-preference approaches with those derived from public decisions, the studies typically
show the inferences from public decisions to yield lower values—not surprising in light of the absence of the hypothetical element in the
public-decision results. Although systematic comparisons with conventional revealed preference approaches are lacking, it is likely that the
valuations of eco-system components calculated from public decisions would be higher, because public decisions do capture whatever
elements of public-regardedness are present among the voters. The valuations based on public decisions have relevance within the paradigm
that gives standing to the community votes as reflecting the policies that the public prefers. Even when a referendum or initiative passes by a
wide margin, which reduces the precision of estimating the value held by the median voter, these outcomes provide strong input to decision
makers regarding publicly held values.

Strengths/Limitations

Willingness to pay (WTP): The results will be most easily interpreted if the initiatives or referenda are: a) as focused as possible on a
single dimension of environmental protection or amenity; b) free of ideological debate; c) confined to easily identifiable government costs
rather than diffused and uncertain costs such as job losses; and d) the wording of the referendum or initiative is both unambiguous and
clarifies the costs to the voters if the measure passes.

Willingness to accept (WTA): The results will be most meaningful if: a) the vote is explicit; b) the expected damage is well specified:
c) property rights are clearly held by the community (i.e., the community has the right to refuse entry) d) the community's gains can be easily
estimated; and e) the transaction costs are low.

The most useful referenda or initiatives would propose direct costs to the voters, typically in the form of taxes, fees, or bonds to
finance actions designed to improve or protect ecosystems. Referenda or initiatives that entail restrictions on development (such as more
stringent emissions or effluent standards) are less useful, because of the uncertainty of the level and incidence of the economic impacts.
Similarly, in order to isolate the values attributed to particular ecosystem benefits, referenda and initiatives that address only one objective,

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such as preserving habitats or reducing air pollution air pollution, are more useful. With multiple objectives, the analysis cannot assign the
willingness to pay to each component. Similarly, if it is clear that a referendum or initiative entails additional partisan political stakes (e.g., if
it is widely viewed as a political test of a government official), the results are less illuminating in terms of the ecosystem values that the voters
hold. The criterion of unambiguous wording is important in light of the findings that the wording of the questions can make a significant
difference in the responses (Cronin 1989, Magleby 1984). However, the problem of misleading wording has been addressed in many
jurisdictions, where election commissions have to approve the wording of both referenda and initiatives. Moreover, the fact that specific
wording can influence responses is obviously not unique to the actual referendum and initiative situations; stated preference approaches, and
surveys in general, face the same wording challenge.

Valuation based on initiative or referendum results would work best when:

•	applied to the same jurisdiction (e.g., if a city is considering another storm control issue, the analysis of that city's referendum
would be most appropriate), but can still be used via benefits transfer;

•	a unitary conservation or environmental benefit is involved;

•	the initiative or referendum outcome was a close vote (this yields stronger inferences about the actual valuation, rather than
floors or ceilings);

•	extraneous issues (such as whether the vote is a "political test" on particular politicians, or the mode of financing is
controversial) are unimportant; and

•	surveys can be accomplished soon after the actual vote.

These approaches attempt to measure the sum total of values of improving or protecting ecosystems and eco-system services;
therefore both means and ends (instrumental and intrinsic) values can be involved. All variants in principle could measure the values
attributed to all types of services, expressed in terms of monetary values per unit of ecosystem improvement or protection. The variants are
flexible in terms of levels of data, detail, and scope, inasmuch as initiatives and referenda decisions have been made at all sub-national levels.

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The valuations can be aggregated across benefits and with other methods, as long as the scale and magnitude of benefits are roughly the same.
While highly complex initiatives and referenda are not good candidates for estimating value, the valuations generated from simpler cases can
be used as inputs for complex applications.

Any EPA decision context calling for monetized valuation could employ these variants, either singly or as cross-checks with
conventional revealed preference or stated preference approaches. Benefit transfer applications will be limited to cases of similar magnitudes
of benefits, because of the likelihood that community decisions are highly sensitive to such magnitudes.

In uses that apply valuations directly to the jurisdiction previously experiencing the initiative or referendum, the scale would be the
same municipality, county, or state. For benefits transfer, the scale should also be the same, given the need for similar magnitude of benefits
and costs mentioned above.

Making valuation estimates directly from referendum or initiative outcomes has two advantages over conventional valuation methods.
Unlike the standard revealed-preference approaches, such as hedonic pricing or the travel-cost method, voting on referenda or initiatives will
reflect as much (or as little) public-regardedness as the voters actually hold toward the objectives involved. Standard revealed-preference
approaches reflect the private utility-maximizing decisions of individuals who purchase homes, spend money to visit parks, etc.; these
decisions do not reflect what individuals want for their communities. Voting affirmatively for referendum- or initiative-proposed public
expenditures does elicit valuing on behalf of the community, insofar as the voters are so disposed. Of course, a voter may vote for or against
a referendum or initiative proposal strictly out of concerns for herself and/or her family, but the outcome does not exclude the existence value
component should it exist.

Unlike the conventional stated preference approaches such as contingent valuation, the analysis based on referendum or initiative
outcomes is not subject to the possible distortions of hypothetically-posed choices. If a voter supports the referendum or initiative proposal,
the vote contributes to the likelihood that the expenditures will actually occur and the costs will actually be borne. Some might argue that the
chance that any one vote will decide the outcome of the referendum or initiative is remote, and therefore the vote is more of a symbolic act
than a trade-off choice. However, there are two important responses to this point. First, whatever the mix of motives of the voters, the

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outcome is the community's decision, and therefore has standing in and of itself. This is the same logic by which we accept elected officials
as legitimate even if we are dubious about the motives or rationality of the voters. Second, even if a voter believes that the chances that his or
her vote will make the difference are negligible, the vote is still an expression of support or opposition to the proposal. There is little reason
to believe that a "yes" vote would reflect just the gratification of voting "yes" (especially in secret balloting) rather than a belief that the
proposal merits support.

Another concern that some would level against inferences based on referenda or initiatives is that these votes are often subject to
intense efforts by interest groups, advocacy groups, and even governments to manipulate public perceptions (Butler & Ranney 1978, Cronin
1989, Magelby 1984). This concern has two aspects: whether the information on which voters base their decisions has been distorted, and
whether the votes are swayed by appeals on one side or the other, especially by the side with the greatest resources (Hadwiger 1992, Lupia
1992, Owens & Wade 1986). The first aspect is more compelling: we certainly would be less willing to accept the validity of an estimate
derived from voting decisions driven by serious misconceptions of the proposed benefits and/or costs. The outcome is still the official
decision of that community, but the justification for using the result as the basis of benefits transfer to other communities would be very weak.
On the other hand, the fact that referenda and initiatives are often subject to intensive campaigns of persuasion may be considered a virtue
rather than a drawback, insofar as it would provide more information on both sides. In addition, the fact that individuals are exposed to
efforts at persuasion is by no means confined to referenda and initiative contests: respondents to contingent valuation surveys have of course
been subjected to many years of promotional activities by environmental groups; people who travel farther to a particularly popular national
park such as Yosemite have been influenced by all sorts of communications extolling its virtues. In short, efforts at value persuasion are
pervasive, and in any event should not be a basis for rejecting the significance of decisions of individuals exposed to those efforts. The
philosophical basis underlying the use of referenda or initiatives, namely that the public's preferences are legitimately shaped by the political
process, and that the public's policy preferences are important beyond how the public values the outcomes that these policies may produce, is
quite different from the "progressivist" position that individuals' values should be determined in isolation of politics (Sagoff 2004: 177-178).

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Another difference in philosophical basis is that the referendum and initiative results reflect intensity of attention to the issue, at least
insofar as those who do not care enough to vote are excluded from the analysis. From the progressivist, technocratic perspective, everyone's
values ought to be incorporated, because the policies ought to maximize utility (i.e., the consequences of public decisions) regardless of
whether specific individuals are mobilized to take action. On the other hand, prominent strains of pluralist democratic theory regard intensity
as a fully legitimate factor in determining policy outcomes (Lowi 1964).

One limitation of estimating values from referendum or initiative outcomes is that it is often difficult for voters to assess the actual
stakes involved. The benefits will often have to be predicted (e.g., how much biodiversity will a reserve really safeguard; how much flooding
will the flood-control system actually prevent?), entailing an amount of uncertainty. The benefits that do occur will often be community-
wide, with some uncertainty as to how much an individual or particular household can take advantage of the benefits. On the cost side, the
burden of a tax increase or bond measure on household expenditures may be very difficult for the typical voter to estimate, and the impacts of
development restrictions may be even more difficult in light of the uncertainty as to which families would ultimately be affected. Insofar as
the costs specified by the referendum or initiative are not easily translatable into household budget terms, the outcome, though it is still "the
community's decision," is less revealing about the values held by the voters.

The outputs of these approaches should be easy to understand and to communicate to the public. It is a significant advantage to be
able to say that the valuation of an ecosystem component has been estimated on the basis of how communities have decided what these
components are worth.

Te\( l$o\ I"7 KiTi'ivmhi iind lni(i;ili\os I sod lo Y;ili(l;ilc ( oniin^cnl Y;ilu;ilion

In addition lo taking the \aliiation deri\ed from the analysis of public decisions as an input in itself, the analysis of public decisions,
particularly referenda and iniliali\es. can be used lo \alidate the results of other \aluation methods Se\eral studies ha\e compiled the results
of initiati\es and or referenda in order to try to \alidate more comenlional \aluation techniques, especially contingent \aluation (kalin &

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1	Malsusaka (llW7). I.isl tK; Shogren (2<)i)2). Murphy, el al (2<)i)3). Schlapler. Roschewiiz. & llanley (2<)i)4). YosslcrcV: kerk\liet (2<)i)3).

2	Yossler. kerk\licl. Polasky tK; (iainiilcliiuna (2<)<>.i) As Arrow, d al (I'W.i) recommend

4	The referendum formal offers one furlhcr achanlaue lor (A As we ha\e argued. external \alidation of elicited lost passi\e use \ allies is

5	usually impossible There are howe\ er rcal-lile referenda. Some of lliem. al least, are decisions lo purchase specific public goods with

6	delined payment mechanisms, eg . an increase in properly taxes The analogy with willingness to pay lor a\ oidance or repair of

7	en\ ironmenlal damage is far from perfect but close enough that the ability of CY-like studies to predict the outcomes of real-world referenda

8	would be useful e\ idence on the \ alidity of the (A' method in general The test we cn\ ision is not an election poll of the usual type. Instead.

9	using the referendum formal and pro\ iding the usual information lo the respondents, a study should ask whether they are willing to pay the

10	a\erage amount implied by the actual referendum The outcome of the (A'-like study should be compared with that of the actual referendum

11	The Panel thinks that studies of this kind should be pursued as a method of \ alidating and perhaps e\en calibrating applications of the (A

12	method

13

14	Does this method incorporate any specific ways of treating uncertainty? Is there any approach unique to this method?

15	There are two distinct sources of uncertainty involved with this approach, depending on which variant is employed and how the

16	outcomes are interpreted. If the referendum or initiative results are used without a follow-up survey, and the results are interpreted as

17	indicating the aggregation of individual valuations, then there is uncertainty as to whether the voters understood the benefits and the payments

18	accurately. If the results are interpreted as the community's preference per se, then the result is accurate in itself, as long as vote miscounting

19	is not an issue.

20	The follow-up survey provides a way to determine whether voters understood the benefits and payments accurately. However, like

21	any survey it also has its own sources of uncertainty: biases in which voters agree to respond to the survey, and untruthfulness in the

22	individual responses. An additional source of potential uncertainty would arise if non-voters are asked to respond to the survey because of

23	error on the part of the survey team. Despite these potential pitfalls, the follow-up survey (equivalent to a contingent valuation study) would

24	serve as a cross-check on the referendum or initiative results.

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Another source of uncertainty in undertaking a benefits transfer of valuation based on referenda or initiatives is that communities
where these efforts are tried may be atypical; for example, it is possible that referenda and initiatives are more likely to be launched in
communities with a stronger commitment to conservation. However, if enough straightforward referenda and initiatives are analyzed and put
into comparable terms, including those that failed to pass, the range of results would provide more robust information than any single result.
Research needs

The research needed to make the results of public decisions through referenda and initiatives most useful for inferring values would
consist of the creation of a data bank of referenda and initiative outcomes, optimally screening out those involving multiple, confounding
elements. Because more than 1,100 referenda on open space issues alone were conducted in the United States between 1997 and 2004
(Banzaf, et al. 2006), the chances are good that a sizable number of referenda will meet the criteria. A preliminary analysis is needed to
determine whether the communities that hold referendum votes are atypical of communities in general (i.e., is there a selection bias among
the referendum-holding communities that would make their valuations atypical of the entire set of communities?). Thus a group of
researchers at Resources for the Future is conducting in-depth analysis of 15 county-level, open-land referenda in Colorado, and also
assessing the other open-land referenda in the rest of the United States (Banzaf, et al. 2006), to determine what kinds of communities hold
referenda and what explains why the majority of referenda pass. The analysis of the valuation of benefits or damage would be
straightforward calculation of the ratios of benefits or costs to the per-household costs, when such ratios can be deduced from simple
referendum or initiative choices. The survey variants would involve considerably more effort of developing the questionnaire, administering
it immediately after a referendum or initiative, and analyzing the additional information, yet the results would provide information on both
median and mean valuation. Once model surveys are developed, they could be used with minor adaptations in different settings. In terms of
resources required to make progress, roughly three researcher-years could produce a credible data base and systematically distill the
information from the voting results that would be useful for policymakers. Using initiative or referendum voting results to cross-validate
other valuation methods can be done at relatively low cost, although the follow-up survey options entail more effort, depending of course on
how elaborate they are.

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Key References

Arrow, K., et al. (1993). Report of the NOAA Panel on Contingent Valuation. Washington, D.C, Government Printing Office.

Ashenfelter, O. and M. Greenstone (2004). "Using Mandated Speed Limits to Measure the Value of a Statistical Life." Journal of Political
Economy 112: S226-S267.

Banzhaf, Spencer, Wallace Oates, James N. Sanchirico, David Simpson, and Randall Walsh. (2006). Voting For Conservation: What Is the
American Electorate Revealing? Resources, Winter (16). Washington, DC: Resources for the Future.
http://ww.rff org/rff/news/features/loader.cfm?url=/commonspot/security/getfile.cfm&pageid=22017.

Butler, David and Austin Ranney, eds. (1978). Referendums. Washington D.C.: American Enterprise Institute.

Cronin, Thomas E. (1989). Direct Democracy: The Politics of Referendum, Initiative and Recall. Cambridge: Harvard University Press.

Deacon, R. and P. Shapiro (1975). "Private preference for collective goods revealed through voting on referenda." American Economic
Review 65: 793.

Hadwiger, David. (1992). "Money, Turnout and Ballot Measure Success in California Cities." Western Political Quarterly 45 (June): 539-547.

Kahn, M. E. and J. G. Matsusaka (1997). "Demand for environmental goods: Evidence from voting patterns on California initiatives." Journal
of Law and Economics 40: 137-173.

List, J. and J. Shogren (2002). "Calibration of Willingness-to-Accept." Journal of Environmental Economics and Management 43: 219-233.

Lowi, Theodore. (1964). "American Business, Public Policy, Case Studies, and Political Theory." World Politics 16:677-715.

Lupia, Arthur. (1992). "Busy Voters, Agenda Control, and the Power of Information." American Political Science Review 86(June): 390-399.

Magleby, David. (1984). Direct Legislation: Voting on Ballot Propositions in the United States. Baltimore, MD: Johns Hopkins University
Press.

Murphy, J. J., P. G. Allen, et al. (2003). A Meta-Analysis of Hypothetical Bias in Stated Preference Valuation, Department of Resource
Economics University of Massachusetts Amherst, MA.

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Owens, John R. and Larry L. Wade. (1986). "Campaign Spending on California Ballot Propositions." Western Political Quarterly 39
(December): 675-689.

Sagoff, Marc (2004). Price, Principle and the Environment. Cambridge: Cambridge University Press.

Schlapfer, F., A. Roschewitz, & N. Hanley (2004). "Validation of stated preferences for public goods: A comparison of contingent valuation

survey response and voting behavior," Ecological Economics, 51: 1-16.

Shabman, L. and K. Stephenson (1996). "Searching for the correct benefit estimate: Empirical evidence for an alternative perspective." Land
Economics 72: 433-49.

Vossler, Christian A. and Joe Kerkvliet (2003). "A criterion validity test of the contingent valuation method: Comparing hypothetical and

actual voting behavior for a public referendum." Journal of Environmental Economics and Management 45(3): 631-49.

Vossler, Christian A., Joe Kerkvliet, Stephen Polasky, and Olesya Gainutdinova. 2003. Externally Validating Contingent Valuation: An
Open-Space Survey and Referendum in Corvallis, Oregon. Journal of Economic Behavior and Organization 51(2): 261-277.

Citizen Valuation Juries
Description of the Method

Another potential process for attempting to measure the social/civic value of changes to ecological systems and services is to assemble
and query a representative group of citizens (a "citizen jury"). The major use of citizen juries to date in environmental decision making has
been to help governments rank options for achieving particular goals, e.g., reducing traffic in an urban area (Kenyon, et al. 2001). Citizen
juries also can be used to measure the value of changes to ecological systems and services along a variety of different metrics. Information
obtained during ranking deliberations, for example, can provide valuable insights for other valuation exercises (Aldred & Jacobs 2000).
Citizen juries also have been combined with choice modeling to determine paired rankings of various ecological characteristics (Alvarez -
Farizo & Hanley 2006).

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Although citizen juries have generally been used to rank governmental options rather than to determine monetary values, citizen juries
can also be asked to determine either a social/civic willingness to pay ("public WTP") or a social/civic willingness to accept ("public WTA")
for any particular ecological change (Blarney, et al. 2000). For public WTP values, citizen valuation juries can be asked to determine the
highest levy, tax, or other form of payment that the government should pay to obtain a particular ecological benefit. For public WTA values,
citizen valuation juries can be asked to determine the highest monetary sum that the government should accept to avoid a particular ecological
loss.

When asked to determine public WTP or public WTA, citizen juries bear both similarities to and differences from initiatives and
referenda and contingent valuations. Like initiatives and referenda, citizen juries provide information on social/civic values, but they measure
stated rather than revealed value, and they incorporate elements of the "deliberative valuation" processes described earlier in this section.
Citizen valuation juries are also similar to contingent valuation surveys except that: a) juries are asked to determine how much the public
should pay or accept in compensation for a specified ecological change (rather than being asked how much they would pay or accept as
individuals); b) valuation juries are often asked to agree on a common value for the ecological change (rather than being asked for individual
values that the expert then aggregates or otherwise combines); c) juries deliberate together as a group before determining value; and d) juries
are provided with more extensive information about the ecological change and can be aided in their deliberations.

Although there is little experience using citizen juries to determine public WTP or public WTA, a number of governmental and
academic experiments have examined the appropriate use of citizen juries to inform various governmental choices more generally. The
process of forming and utilizing citizen juries has varied widely. In the typical situation, a small group of citizens, typically ranging from a
cross-section of 12 to 20 persons, has been drawn from the relevant population. Approaches have differed as to how best to choose the jurors.
Given the small size of citizen juries, there is an inevitable tension between choosing jurors to reflect the demographic characteristics of the
relevant population as a whole and choosing jurors that represent the interests of major stakeholders. Although larger juries would reduce
some of the tensions involved in juror selection, larger juries are likely to find it more difficult to reach agreement within a realistic time
frame. Most citizen juries to date have been chosen using random sampling or stratified random sampling (Blarney, et al. 2000).

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Once a citizen jury is chosen, the jury then meets and deliberates over a multi-day period, during which it hears and questions expert
witnesses, deliberates in small and large groups, and agrees on a final recommendation to the sponsoring governmental body. These group
deliberations allow jurors to hear alternative perspectives, test ideas, and carefully work through the valuation exercise. Several different
techniques are used to provide information to the jurors. In some cases, the government or an expert facilitator chooses what information to
provide to jurors, while in other cases, relevant interest groups make individual presentations to the jury. Jurors also can be permitted to
request information and pose questions directly to expert witnesses (Blarney, et al. 2000). Two factors should guide choices among the
processes for providing information to the jurors: a) ensuring that jurors have all the information that they believe is valuable to their
valuation exercise; and b) ensuring that the information is balanced and not biased toward any particular result. Another important choice in
designing a citizen jury is the process by which the jury will make decisions. In most cases, juries are asked to arrive at a group decision.
Decision making rules in this context include a simple majority vote of the jury, consensus (where a majority favors the valuation and no
juror opposes it), and unanimous agreement. Citizen juries also do not need to produce a collective value. In some experiments, for example,
juries deliberate as a group, but members of the jury then report their valuations on an individual basis (Alvarez-Farizo & Hanley 2006).
Researchers can then combine individual valuations into an overall evaluation. Measures of central tendency (means or mediums of the
valuations provided by the individual jurors) can be used to develop a valuation measure in this context.

Experiments indicate that citizen juries often produce significantly different valuation results from economic or socio-psychological
surveys. The additional information available to jury members, the opportunity to spend time thinking about the appropriate valuation, and
the stress on collective rather than individual values all appear to generate significant changes in valuation (Alvarez-Farizo & Hanley 2006).
The jury's valuation of particular ecological improvements, however, can either increase or decrease compared to the results obtained through
economic surveys (Alvarez-Farizo & Hanley 2006).

Because contingent valuation methodology and other traditional economic measurement approaches seek a very different valuation
than citizen valuation juries, juries should not be seen as a substitute for the traditional approaches. Governmental agencies should employ
citizen valuation juries as a supplement to traditional economic valuation approaches. When deciding whether to pursue particular

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regulations or other governmental actions, agencies should consider estimates of both private and public value, along with the strengths and
weaknesses of each approach.

EPA might also consider using some elements of the citizen jury approach to improve other valuation methods. Some researchers
have investigated other group-based approaches out of concern, for example, about whether contingent valuation surveys provide sufficient
time and information for survey respondents to generate reliable estimates of the value of often complex ecological changes. Under the
"Market Stall" ("MS") approach, for instance, researchers meet with survey subjects in two one-hour meetings, separated by a week, and
encourage the participants to discuss their valuations with household members and friends between the two sessions. Unlike citizen valuation
juries, the MS approach asks survey subjects for their personal valuations, based on individual preferences and incomes, rather than
social/civic valuation. Respondents are asked for their personal valuations in a confidential written survey at the end of the second meeting.
In Macmillan, et al. (2002), the WTP measures obtained through the MS approach were significantly lower than the WTP measures generated
from CV interviews, which is consistent with other studies that show a decline in WTP when survey subjects are provided additional time to
consider their answers (Whittington, et al. 1992).

Text Box 18: A Valuation Exercise Illustrating Use of Citizen Juries

In one experiment, a citizen jury was used to examine the economic value of the control of a particular exotic weed, Bitou Bush
(Chrysanthemoides monilifera L. Norl. ssp. rotundata), in an Australian national park (James & Blarney 2000). A jury of 14 was
selected, using a two-phase telephone survey, in order to be representative of the New South Wales population on the basis of gender,
age, place of residence, rating of the environment in relation to other social issues, occupation, income, income source, and education.
The jury met for three days during which they heard and questioned seven expert witnesses. Prior to the hearings, jurors received
training in note taking and questioning of witnesses, in order to maximize their ability to use the information provided.

In one of the charges, the jury was given two options: (Option #1) the then-current situation in which weeds were controlled on 3000
hectares per year, and (Option #4) an alternative management regime in which weed control would be expanded to 9600 hectares per
year. The jury was then given the following charge: "How high would a park management levy have to be, before the jury would

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recommend Option I rather than Option 4 In other words, how hiuh would the le\\ ha\e to he before the puhlic would he no
better off under Option 4 than Option I"'" The jury first decided that a prouressi\e le\y. calculated as a percent of gross income, was
most appropriate After discussing two proposed le\ ies (<) l"n and <> 25"<>). the jury \ oted eight to two in la\ or of a le\ y of <) l"n In a
siiia ey following the jury exercise, jurors reported that they found the \ aluation exercise to he both interesting and worthwhile

Relation of Method to the C-VPESS Expanded and Integrated Framework

Citizen juries are potentially useful both to identify socially important assessment endpoints and to attach a value, monetary or socio-
psychological, to changes in the assessment endpoints. Use of this method relates to steps 3 and 5 of the C-VPESS proposed valuation
process (Figure 2).

Because citizen juries consist of representative members of the public, citizen juries also expand the role that the public plays in
valuations of changes in ecological systems and services. Members of citizen juries actively evaluate information regarding changes, are
permitted to ask questions of experts, and consciously deliberate over the appropriate social/civic value of the change.

Status as a Method

As discussed earlier, citizen juries have been used primarily to help governments rank options for achieving particular goals. Only a
few efforts have been made to date to use citizen juries to generate monetary or other estimates of the social/civic value of environmental
changes. Use of citizen juries for direct valuation of changes to ecological systems and services, therefore, should be considered experimental
for the moment and should not be used to make significant governmental decisions until further research has been conducted on both the
efficacy of the process and the appropriate jury processes. Given the potential use of citizen juries to evaluate social/civic values, however,
this is an area in which research can be valuably focused. EPA may wish to use citizen juries on an experimental basis, moreover, to provide
a comparison to valuations obtained through traditional economic valuation methods.

Strengths/Limitations

One of the major strengths of a citizen valuation jury is that, like referenda and initiatives, the citizen valuation jury incorporates
public-regardedness. Jurors are asked to provide a valuation based on the perceived impact of an ecological change on the entire community

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rather than on his or her individual preferences alone. Citizen valuation juries thus incorporate a broader concept of value than standard
contingent valuation approaches and place the jurors in a position similar to that of the governmental decision makers who are being advised.

Citizen valuation juries avoid a number of potential concerns regarding referenda and initiatives as a source of social/civic valuation
information. First, the jury process ensures that juries receive more information regarding the ecological change than most voters receive
prior to voting on an initiative or referendum. Second, because the jury evaluation process can be carefully structured, citizen evaluation
juries are less subject to undue influence from political interest groups than are votes on referenda and initiatives. Finally, there are a limited
number of referenda and initiatives from which valuations can be derived, while citizen valuation juries can be asked to assess a valuation for
any ecological change. Unlike referenda and initiatives, however, citizen juries do not have standing as actual, official decision making
bodies for their communities.

Citizen valuation juries build on a well-established legal institution in the United States - the criminal and civil jury system. The legal
system uses juries to decide whether to initiate criminal prosecutions, determine guilt and innocence in criminal cases, decide between life
and death in capital cases, and assess damages in often complex civil cases. Most adult members of the public have served as jurors,
understand the importance of the role they assume, and act deliberately and responsibly.

Citizen valuation juries suffer from the hypothetical character of all stated-value methods of valuation. Because the juries do not
themselves determine governmental policy, the juries may not reveal what they actually believe to be the social/civic value of an ecological
change. The hypothetical character of jury valuations could be eliminated by providing that the valuations will directly determine whether
particular governmental actions will be taken, but the government is unlikely to want to (or be legally able to) delegate its decision making
powers to citizen juries. Despite concerns over hypothetical inquiries, experiments with citizen juries indicate that jurors approach their
valuation task in a responsible fashion and reach well-thought-out conclusions (Aldred & Jacobs 2000).

Citizen juries also raise a number of other unique concerns. Some economists, for example, have worried that group dynamics and
"norms" might reduce the reliability of jury decisions. Some jurors, for example, might not wish to be perceived as disagreeing with others,
while some jurors may be able to dominate the discussion and result. Some jury experiments, however, have suggested that the design of the

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jury process can avoid such jury dynamics (Macmillan, et al. 2002). Trained facilitators may be able to overcome any structural pathology
that might otherwise arise and should be involved in any valuation exercise involving citizen juries.

As discussed earlier, the choice of jurors also poses difficulties. Because of the small size of typical citizen juries, a demographic
cross-section of the public may not adequately represent all interest groups. Choosing representatives of different interest groups to serve on
citizen juries, however, may yield a jury that does not adequately represent demographics. Small citizen juries, moreover, will inevitably fail
to fully represent the public as a whole. In order to ensure that jurors are other-regarded, experiments suggest that the government should
choose a jury that is as demographically representative as possible (typically through stratified random sampling), so that the jury is at least
symbolically representative, and then instruct the jury to adopt an impartial stance in its deliberations (Brown, et al. 1995, Blarney, et al.
2000).

Treatment of Uncertainty

The use of citizen juries to value changes in ecological systems and services raises many of the same uncertainties as traditional
methods of economic or socio-psychological valuation. The small size of citizen juries, however, raises an additional uncertainty factor.
Research Needs

Because there is little experience with the use of citizen juries to directly value changes in ecological systems and services, further
research is needed on a variety of topics before EPA should consider adopting the approach to develop social/civic valuations for decision
making purposes on other than an experimental basis. Key questions include:

•	Do citizen valuation juries arrive at different valuations than individual respondents to CV surveys? If so, how and why do the
valuations differ?

•	How stable are valuations provided by citizen juries? How much variation exists among the valuations produced by different
citizen juries?

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•	How do jury selection processes affect the valuations of the jury? What methods exist to overcome the inevitable bias arising
from the small size of citizen juries?

•	How should information be provided to citizen valuation juries? What are the advantages and disadvantages of highly
structuring the information that is provided to a jury, versus permitting the jury to determine the information that it receives?

•	How do decision making rules (e.g., consensus versus unanimity) affect valuations? What are relevant considerations in
choosing among the different decision making rules?

Key References

Aldred, J. and M. Jacobs (2000). "Citizens and Wetlands: Evaluating the Ely Citizens' Jury." Ecological Economics 34: 217-232.
Alvarez-Farizo, B. and N. Hanley (2006). "Improving the Process of Valuing Non-Market Benefits: Combining Citizens' Juries with Choice

Modelling." Land Economics 82(3): 465-478.

Blarney, R.K., et al. (2000). Citizens' Juries and Environmental Value Assessment. Canberra, Australia, Research School of Social Sciences,

Australian National University.

Brown, T.C., et al. (1995). "The Values Jury to Aid Natural Resource Decisions." Land Economics 71(2): 250-260.

Gregory, R. and K. Wellman (2001). "Bringing Stakeholder Values into Environmental Policy Choices: A Community-Based Estuary Case

Study." Ecological Economics 39: 37-52.

Kenyon, W. and N. Hanley (2001). "Economic and Participatory Approaches to Environmental Evaluation." Glasgow, U.K., Economics

Department, University of Glasgow.

Kenyon, W. and C. Nevin (2001). "The Use of Economic and Participatory Approaches to Assess Forest Development: A Case Study in the
Ettrick Valley." Forest Policy and Economics 3(1): 69-80.

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1	Macmillan, D.C., et al. (2002). "Valuing the Non-Market Benefits of Wild Goose Conservation: A Comparison of Interview and Group-

2	Based Approaches." Ecological Economics 43: 49-59.

3	McDaniels, T.L., et al. (2003). "Decision Structuring to Alleviate Embedding in Environmental Valuation." Ecological Economics 46: 33-

4	46.

5	O'Neill, J. and C.L. Spash (2000). "Appendix: Policy Research Brief: Conceptions of Value in Environmental Decision-Making."

6	Environmental Values 9: 521-536.

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DELIBERATIVE PROCESSES

Mediated Modeling

Brief description of the method

Computer models of complex systems are frequently used to support decisions concerning environmental problems. To effectively use
these models, (i.e., to foster consensus about the appropriateness of their assumptions and results, and thus to promote a high degree of
compliance with the policies derived from the models) it is not enough for groups of academic experts to build and run the models. What is
required is a different role for modeling - as a tool in building a broad consensus across academic disciplines as well as between science and
policy.

Mediated modeling is a process of involving stakeholders (parties interested in or affected by the decisions the model addresses) as
active participants in all stages of the modeling, from initial problem scoping to model development, implementation, and use (Costanza and
Ruth 1998, van den Belt 2004). Integrated modeling of large systems, from individual companies to industries to entire economies or from
watersheds to continental scale systems and ultimately to the global scale, requires input from a very broad range of people. We need to see the
modeling process as one that involves not only the technical aspects, but also the sociological aspects involved with using the process to help
build consensus about the way the system works and which management options are most effective. This consensus needs to extend across the
relevant academic disciplines, the science and policy communities, and the public. Appropriately designed and appropriately used mediated
modeling exercises can help to bring these communities together. The process of mediated modeling can help to build mutual understanding,
solicit input from a broad range of stakeholder groups, and maintain a substantive dialogue between members of these groups. Mediated
modeling and consensus building are also essential components in the process of adaptive management (Gunderson, Holling, and Light 1995,
van den Belt 2004).

Example of how the method could be used as part of the C-VPESS expanded and integrated framework

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As described, the method is fairly general and could be used to assess any value that a group of stakeholders could identify and build
into a model. Any decision context that requires the estimation of the values of ecosystem goods or services could employ this method,
although to the committee's knowledge no EPA decisions have as yet employed this technique. The method covers all elements of the diagram
representing the C-VPESS framework for valuation after the initial identification of EPA needs, and could be used in conjunction with the full
range of decision models. Prior applications have been at a broad range of scales, from watersheds or specific ecosystems to large regions and
the global scale. The method is in principle broadly applicable to the full range of time and space scales.

•	The method is inherently dynamic.

•	The results can be aggregated to get a single benefits number as needed.

•	Participants in the mediated modeling process gain deep understanding of the process and products, if the process is done well. Those
who have not participated can easily view and understand the results if they invest the effort. Usually the results can (with some
additional effort) be made accessible to a broad audience.

•	Since the method explicitly discusses and incorporates subjective or "framing" issues, it is at least open and transparent to users.

Status as a method

As mentioned above, mediated models can contain explicit valuation components. In fact, if the goal of the modeling exercise is to
consider trade-offs, then valuation of some kind becomes an essential ingredient. How these trade-offs and valuations are incorporated into the
model varies, of course, from exercise to exercise. Perhaps the best way to describe this process is with an example. The South African fynbos
ecological/economic model described by Higgins, et al. (1997) is an illustrative example.

The area of study for this example was the Cape Floristic Region—one of the world's smallest and, for its size, richest floral kingdoms.
This tiny area, occupying a mere 90,000 km2, supports 8,500 plant species, 68 percent of which are endemic (193 endemic genera and six

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endemic families [Bond and Goldblatt 1984]). Because of the many threats to this region's spectacular flora, it has earned the distinction of
being the world's "hottest" hot-spot of biodiversity (Myers 1990).

The predominant vegetation in the Cape Floristic Region is fynbos, a hard-leafed and fire-prone shrubland which grows on the highly
infertile soils associated with the ancient, quartzitic mountains (mountain fynbos) and the wind-blown sands of the coastal margin (lowland
fynbos) (Cowling 1992). Owing to the prevalent climate of cool, wet winters and warm, dry summers, fynbos is superficially similar to
California chaparral and other Mediterranean climate shrublands of the world (Hobbs, Richardson, and Davis 1995). Fynbos landscapes are
extremely rich in plant species (the Cape Peninsula has 2,554 species in 470 km2) and plant species endemism ranks amongst the highest in the
world (Cowling 1992).

In order to adequately manage these ecosystems, several questions had to be answered including: what services do these species-rich
fynbos ecosystems provide and what is their value to society? A two-week workshop was held at the University of Cape Town (UCT) with a
group of faculty and students from different disciplines along with parks managers, business people, and environmentalists. The primary goal
of the workshop was to produce a series of consensus-based research papers that critically assessed the practical and theoretical issues
surrounding ecosystem valuation as well as assessing the value of services derived by local and regional communities from fynbos systems.

To achieve these goals, an 'atelier' (or combined workshop/short course) approach was used to form multidisciplinary, multicultural
teams, breaking down the traditional hierarchical approach to problem solving. Open space (Rao 1994) techniques were used to identify critical
questions and allow participants to form working groups to tackle those questions. Open space meetings are loosely organized efforts that give
all participants an opportunity to raise issues and participate in finding solutions.

The working groups of this workshop met several times during the first week of the course and almost continuously during the second
week. The groups convened together periodically to hear updates of group projects and to offer feedback to other groups. Some group
members floated to other groups at times to offer specific knowledge or technical advice.

Despite some initial misgivings on the part of the group, the structure of the course was remarkably successful, and by the end of the
two weeks, seven working groups had worked feverishly to draft papers. These papers were eventually published as a special issue of

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Ecological Economics (Cowling and Costanza 1997). One group focused on producing an initial scoping (or mediated) model of the fynbos.
This modeling group produced perhaps the most developed and easiest-to-implement product from the workshop: a general dynamic model
integrating ecological and economic processes in fynbos ecosystems (Higgins, et al. 1997). The model was developed in STELLA and
designed to assess potential values of ecosystem services given ecosystem controls, management options, and feedbacks within and between the
ecosystem and human sectors. The model helped to address questions about how the ecosystem services provided by the fynbos ecosystem at
both a local and international scale are influenced by alien invasion and management strategies. The model consists of five interactive sub-
models: a) hydrology; b) fire; c) plants; d) management; and (e) economic valuation. Parameter estimates for each sub-model were either
derived from the published literature or established by workshop participants and consultants (they are described in detail in Higgins, et al.
1997). The plant sub-model included both native and alien plants. Simulation of the model produced a realistic description of alien plant
invasions and their impacts on river flow and runoff.

This model drew in part on the findings of the other working groups and incorporates a broad range of research by workshop
participants. Benefits and costs of management scenarios were addressed by estimating values for harvested products, tourism, water yield, and
biodiversity. Costs included direct management costs and indirect costs. The model showed that the ecosystem services derived from the
Western Cape Mountains are far more valuable when vegetated by fynbos than by alien trees (a result consistent with other studies in North
America and the Canary Islands). The difference in water production alone was sufficient to favor spending significant amounts of money to
maintain fynbos in mountain catchments.

The model was designed to be user-friendly and interactive, allowing the user to set such features as area of alien clearing, fire
management strategy, levels of wildflower harvesting, and park visitation rates. The model has proven to be a valuable tool in demonstrating to
decision makers the benefits of investing now in tackling the alien plant problem, since delays have serious cost implications. Parks managers
have implemented many of the recommendations flowing from the model.

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There are several other case studies in the literature of various applications of mediated modeling to environmental decision making,
including valuation. Van den Belt (2004) is the best recent summary and synthesis. Some additional examples of mediated modeling projects
where ecosystem service values were integrated are:

•	Participatory Energy Planning in Vermont, Department of Public Service in Vermont,
http://www.publicservice.vermont.gov/planning/mediatedmodeling.html

•	Mediated Modeling of the impacts of Enhanced UV-B Radiation on Ecosystem Services (van den Belt, et al. 2006)

•	Ria Formosa Coastal Wetlands (a case study in van den Belt 2004)

•	Upper Fox River Basin (a case study in van den Belt 2004)

•	A consensus-based simulation model for management of the Patagonian coastal zone (van den Belt, et al. 1998)

Models can be downloaded from: www.mediated-modeling.com
Strengths/Limitations

Resources needed to implement the method vary from application to application. The method can deal with a broad range of available
data and resources, probably better that most other methods, since the model can adapt to the resources available across different levels of data,
detail, scope and complexity. As a rule of thumb, one can produce a credible mediated model in 30-40 hours of workshops, requiring about
300-400 hours of organizing/modeling. Cost: about $40,000 - $100,000 depending on side activities.

The most serious obstacle seems to be the fact that this method is very different from the top-down approach most frequently used in
government. It requires that consensus-building be put at the center of the process, which can be very scary for institutions accustomed to
controlling the outcome of decision processes. An institutional mandate is important, however, to motivate various stakeholders to volunteer
their time, knowledge, and energy to a mediated modeling process. The final outcome of this process cannot be predetermined.

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Treatment of Uncertainty

In terms of uncertainty, there are all the usual sources, but the difference is that the stakeholders are exposed to these sources as they go,
and learn to understand and accommodate them as part of the process. The method is compatible with formal or informal characterizing of
uncertainty, producing probability distributions in addition to point estimates.

Research needs

No research has yet been done on whether application of the process to exactly the same problem by multiple independent groups would
yield "consistent and invariant" results. One would expect general consistency, but some variation between applications. This is an area for
further research.

To evaluate the impact of a mediated modeling process, surveys have been used before and after a process in the past and this research
would deepen the understanding about exactly what elements of a mediated modeling process contribute to the success or failure of these
processes.

Key References

Bond, P. and Goldblatt, P. 1984. Plants of the Cape Flora. Journal of South African Botany Supp. 13:1-455.

Checkland, P. 1989. Soft Systems Methodology, in J. Rosenhead (ed.) Rational Analysis for a Problematic World, John Wiley and Sons,
Chichester, England.

Costanza, R. 1987. Simulation Modeling on the Macintosh Using STELLA, Bioscience, Vol. 37, pp. 129 - 132.

Costanza, R., F. H. Sklar, and M. L. White. 1990. Modeling Coastal Landscape Dynamics. Bioscience 40:91-107

Costanza, R. and M. Ruth. 1998. Using dynamic modeling to scope environmental problems and build consensus. Environmental Management
22:183-195.

Costanza, R., A. Voinov, R. Boumans, T. Maxwell, F. Villa, L. Wainger, and H. Voinov. 2002. Integrated ecological economic modeling of the
Patuxent River watershed, Maryland. Ecological Monographs 72:203-231.

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Cowling, R.M. (ed.). 1992. The ecology of fynbos. Nutrients, fire and diversity. Oxford University Press, Cape Town.

Cowling, R. and R. Costanza (eds). 1997. Valuation and Management of Fynbos Ecosystems. Special section of Ecological Economics vol 22,
pp 103-155.

Ford, A. 1999. Modeling the Environment: An Introduction to System Dynamics Models of Environmental Systems. Island Press, Washington,
DC

Gunderson, L. C. S. Holling, and S. Light (eds). 1995. Barriers and bridges to the renewal of ecosystems and institutions. Columbia University
Press, New York. 593 pp.

Hannon, B. and M. Ruth. 1994. Dynamic Modeling, Springer-Verlag, New York.

Hannon, B. and M. Ruth. 1997. Modeling Dynamic Biological Systems, Springer-Verlag, New York.

Higgins, S. I., J. K. Turpie, R. Costanza, R. M. Cowling, D. C. le Maitre, C. Marais, and G. Midgley. 1997. An ecological economic simulation
model of mountain fynbos ecosystems: dynamics, valuation, and management. Ecological Economics 22:155-169.

Hobbs, R.J., Richardson, D.M. and Davis, G.W. 1995. Mediterranean-type ecosystems: opportunities and constraints for studying the function
of biodiversity. In: Mediterranean-type ecosystems. The function of biodiversity. G.W. Davis and D.M. Richardson (eds), pp 1-42.
Springer, Berlin.

Kahnemann, D. and A. Tversky. 1974. Judgment Under Uncertainty, Science, Vol. 185, pp. 1124 - 1131.

Kahnemann, D., P. Slovic, and A. Tversky. 1982. Judgment Under Uncertainty: Heuristics and Biasis, Cambridge University Press,
Cambridge.

Lyneis, J.M. 1980. Corporate Planning and Policy Design: A System Dynamics Approach, Pugh-Roberts Associates, Cambridge,
Massachusetts.

Morecroft, J.D.W. 1994. Executive Knowledge, Models, and Learning, in J.D.W. Morecroft, and J.D. Sterman (eds.) Modeling for Learning
Organizations, Productivity Press, Portland, Oregon, pp. 3 - 28.

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Morecroft, J.D.W., D.C. Lane and P.S. Viita. 1991. Modelling Growth Strategy in a Biotechnology Startup Firm, System Dynamics Review,
No. 7, pp. 93-116.

Morecroft, and J.D. Sterman (eds.) 1994. Modeling for Learning Organizations, Productivity Press, Portland, Oregon

Myers, N. 1990. The biodiversity challenge: expanded hot-spots analysis. The Environmentalist 10: 243-255.

Peterson, S. 1994. Software for Model Building and Simulation: An Illustration of Design Philosophy, in J.D.W. Morecroft, and J.D. Sterman
(eds.) Modeling for Learning Organizations, Productivity Press, Portland, Oregon, pp. 291 - 300.

Phillips, L.D. 1990. Decision Analysis for Group Decision Support, in C. Eden and J. Radford (eds.) Tackling Strategic Problems: The Role
of Group Decision Support, Sage Publishers, London.

Rao, S.S. 1994. Welcome to open space. Training (April): 52-55.

Richmond, B. and S. Peterson. 1994. STELLA II Documentation, High Performance Systems, Inc., Hanover, New Hampshire.

Roberts, E.B. 1978. Managerial Applications of System Dynamics, Productivity Press, Portland, Oregon.

Rosenhead, J. (ed.) 1989. Rational Analysis of a Problematic World, John Wiley and Sons, Chichester, England.

Senge, P.M. 1990. The Fifth Discipline, Doubleday, New York.

Simon, H.A. 1956. Administrative Behavior, Wiley and Sons, New York.

Simon, H.A. 1979. Rational Decision-Making in Business Organizations, American Economic Review, Vol. 69, pp. 493 - 513.

Van den Belt, M. 2004. Mediated Modeling: a systems dynamics approach to environmental consensus building. Island Press, Washington, DC.

Van den Belt, Maijan, Oscar Bianciotto, Robert Costanza, Serge Demers, Susana Diaz, Gustavo Ferryra, Evamaria Koch, Fernando Momo,
Maria Vernet. 2006. Mediated Modeling of the impacts of Enhanced UV-B Radiation on Ecosystem Services. Photochemistry and
Photobiology, 82: 865-877

Van den Belt, M.J., L.Deutch, A.Jansson, 1998. A consensus-based simulation model for management in the Patagonia coastal zone, Ecological
Modeling, 110:79-103Vennix, J. A. M. 1996. Group Model Building : Facilitating Team Learning Using System Dynamics. Wiley, NY.

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1	Vennix, J. A.M. and J.W. Gubbels. 1994. Knowledge Elicitation in Conceptual Model Building: A Case Study in Modeling a Regional Dutch

2	Health Care System, in J.D.W. Morecroft, and J.D. Sterman (eds.) Modeling for Learning Organizations, Productivity Press, Portland,

3	Oregon, pp. 121 - 146.

4	Westenholme, E.F. 1990. System Inquiry: A System Dynamics Approach, John Wiley and Sons, Chichester, England.

5	Westenholme, E.F. 1994. A Systematic Approach to Model Creation, in J.D.W. Morecroft, and J.D. Sterman (eds.) Modeling for Learning

6	Organizations, Productivity Press, Portland, Oregon, pp. 175 - 194.

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Valuation by Decision Aiding

Decision aiding approaches provide a method for valuing protection of ecological systems and services in terms of multiple attributes.
These approaches are deliberative in nature, rely upon insights drawn from the discipline of decision analysis, and are based on research and
practical findings from applications of decision-aiding approaches (Arvai & Gregory 2003a, Arvai, et al. 2001, Gregory, et al. 2001a, Gregory,
et al. 2001b). Decision-aiding approaches consider value to be a product of a two-step process.

The first part of the process assists people in determining value based on a careful and comprehensive analysis of the suite of attributes
that characterize ecological systems and services. For example, people may determine the value of an estuary based on multiple, ecologically-
based attributes such as the degree to which it provides nutrient exchange, the re-supply of dissolved oxygen to near-shore habitat, or nursery
habitat for anadromous fish species. Similarly, the value of the estuary will also be affected by a wide range of attributes that reflect economic
or social interests, such as the degree to which it provides access to commercially important species, opportunities for recreation, and lanes for
shipping traffic. Decision-aiding approaches consider both types of attributes.

The second aspect of these decision-aiding approaches focuses on helping people form judgments about the value of ecological systems
and services by way of a comparative framework. From a prospective standpoint, decision-aiding approaches help people to evaluate
competing alternatives, determining, for example, which option in a range of environmental, risk, or resource management options is most
likely to lead to a preferred suite of outcomes. In other words, this approach helps people determine which option, in a set, is most valuable
(i.e., is Option A in a set of alternatives better or more valuable to decision makers than Option B?). The value of ecological systems and
services can also be determined retrospectively by comparing attributes associated with ecosystem health or the provision of ecological services
that have been realized today with those that were realized at some point in the past (i.e., is the system being evaluated "better off—or more
valuable—today, at Time 2, than it was in the past, at Time 1?). Alternatively, value can be determined in a spatial comparison by evaluating
the attributes associated with ecosystem health or the provision of ecological services in an area of interest relative to those that have been
realized elsewhere (i.e., is System A more valuable than System B?).

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It is important to note that valuation by decision aiding does not provide an estimate of how valuable ecological systems and services
are. For example, this method cannot provide a specific estimate, which would state that a system today is X times more valuable than it was in
the past, or that System A is Y times more valuable than System B. The concept, which is adapted from a framework for making choices
among options, is ideally suited to providing a relative ranking of value or importance such as when EPA may wish to prioritize systems for
management action.

In the important first step of valuation by decision-aiding process, one or more analysts facilitate the characterization of the ecological
system (or systems) that is to be the focus of analysis. This step in this process entails identifying the relevant attributes of the ecological
system, that is, all aspects of a system that are of interest or concern to people. The goal at this stage is to develop an explicit, comprehensive
picture of all factors that contribute significantly to the overall value of the system in question. Diverse groups of stakeholders and relevant
experts should be consulted to identify the attributes that will ultimately guide the analysis. These stakeholders are defined in an operational
sense as groups of people who, for any reason—e.g., place of residence, occupation, favored activities—have legitimate concerns or opinions
regarding the health of an environmental system. Careful selection of stakeholder groups ensures that the full range of views is adequately
covered. For example, the representatives of an environmental advocacy organization might be expected to present a somewhat different list of
attributes than would representatives of industry or government, but the views of each group are likely to encompass those of many other
citizens.

In addition to consulting the broad spectrum of interested or affected stakeholders, an analyst should also consult with technical experts (e.g.,
ecologists, toxicologists, economists, behavioral scientists, etc.) as part of an interdisciplinary, analytic-deliberative process (Environmental
Protection Agency 2000, National Research Council 1996) designed to identify both the relevant attributes of the system in question as well as
the specific means by which each attribute can be measured (see Text Box 19: Types of Attributes).

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Tc\l l$o\ IV l \|K's ol' AUribuk's

hexious work (keeney ll^2. keeney tK; (iregory 2'»(>5) has led lo an operational typology of attribute to inform their selection in a
gi\en \aliiation context (ienerally speaking. attri hutes that help lo define the dilVerent aspects of a system fall into one of three categories

•	Natural attri hutes these are direct measure conditions that exist in a system I'or example, if one attribute of an en\ ironmental
system being e\aluated is the economic \ alLie of a commercially important species (e g . fish or trees), then the specific \alue of
this attribute can be expressed directly in dollars. Likewise, if an attribute of a system is the number of indi\ iduals of a key
indicator species li\ ing in it. then a straightforward count of these indi\ iduals represents another direct measure of \ al Lie

•	Proxy attributes - these, by contrast, are used when it is not possible to directly measure an attribute of interest lor example, if
one attribute of an en\ ironmental system is the recreational opportunities that it pro\ ides to tourists, economists may by
proxy estimate, using the tra\ el cost method, the recreational \ al Lie of the resource Similarly, a particular miklfhil may be

\ allied from an ecological standpoint because of the migratory shorebirds that it attracts I lowe\er. it is frequently the case that
accurate, direct counts of shorebirds. which would a be natural attribute, are impossible to achie\e In these cases, an analyst
may rely upon the amount of habitat that is a\ ailable as a reasonable proxy for the number of shorebirds that may use the
mudllat o\er the course of a season

•	Constructed attributes these are most often used when neither a direct, natural attribute nor a reasonable proxy attribute exists
Proxy attributes are typically used to operationali/.e objecti\ es that are psychophysical in nature (eg. the objecti\ e to impro\ e
the aesthetic quality of a shoreline) Scales that may be administered during sur\cys often need lo be constructed eg . by
psychologists or sociologists as a means of characterizing these attributes

In the second step of this process, data or information about each of the identified attributes must be collected by those familiar with
how to conduct the individual valuation methods (e.g., ecological, economic, psychosocial, etc.) discussed elsewhere in this report. This
information must be collected at the site of primary interest as well as at other sites that will provide the basis for comparison. Alternatively,
contemporary data at a site of interest must be collected and compared with archived information about previous conditions described by the
same attributes at the site.

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All this information, which describes both the attributes of an ecological system and
specific information to be used as the basis for making comparisons (e.g., data describing
conditions at another site or the same site at an earlier time), can be displayed visually in a
matrix (Table 9). It is unlikely, except in very rare circumstances, that comparisons made
apparent by this matrix will reveal improvements (or, on the other hand, declines) in the
values associated with all of the attributes; in most cases, the comparison will reveal that
improvements have been realized across some attributes while declines have occurred
across others. In the hypothetical estuary described on the previous page, for example, it is
not uncommon for improvements in the system's capacity for nutrient exchange to come at
the expense of opportunities for recreation or industry.

These differences necessitate the need for trade-offs—the third step in a valuation by decision-aiding process—across the attributes to
determine if, on aggregate (1) a site, System A, is more valuable than another, System B, or (2) the system being evaluated, again System A, is
more valuable today than it was in the past (Table 1). A detailed overview of specific methods for addressing these trade-offs, such as swing-
weights (e.g., see Clemen 1996) or even swaps (e.g., see Keeney 1992), are beyond the scope of this discussion. However, these and other
methods can be used by individuals or in deliberating groups to place weights on the various attributes, and in turn, to use these weights to
develop an understanding of the overall, multi-attribute value associated with an environmental system of interest. In other words, despite the
fact that conditions described by certain attributes may have improved while others may have declined, formal trade-off analysis across these
attributes can help individuals or groups decide if conditions on the whole at a site are better or worse—i.e., have higher or lower value—
relative to the reference condition.

Thus far, this discussion has not focused on the situation where people may wish to establish the multi-attribute value of an
environmental system absent a comparative framework for trade-off analysis. Carrying out this kind of assessment is possible and requires that,



Option



Site



Time



A

B



A

B



1

2

Attribute
1

















Attribute

2





0





0





Attribute

3





r





r





Attribute
n

















Table 9: Comparative Matrices of Attributes for
Three Hypothetical Decision-Aiding Valuation
Scenarios

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in lieu of a comparison, individuals or deliberative groups translate the information obtained for each attribute (e.g., inputs in dollars for
attributes that require monetization, constructed scales for attributes measured using psychosocial methods, etc.) into common terms.

Suppose, for example, that EPA wished to construct a value for the damage resulting from a specific pollutant accidentally spilled into a
waterway. Technical experts working alongside stakeholders could be engaged in a process to both identify the relevant attributes of the
system and provide information describing the conditions in the waterway as they relate to these attributes both before and after the insult to the
system. For example, the physical event of the death of a large number of fish might imply not only an ecological loss, but also aesthetic (e.g.,
when the dead fish wash up on shore) and economic (e.g., the loss of commercial fishing jobs and profits) losses. Clearly, a host of other
attributes would also need to be considered.

After the attributes have been identified and the quantitative information that describes them collected, deliberation and argument can be
organized with the intent of deriving a single metric (e.g., dollars or units of ecological productivity) that can be used to capture information
about all of the attributes. For example, the techniques of multi-attribute utility theory (Keeney & Raiffa 1993) can be used to construct a
single "value" that encompasses the diverse array of attributes (Gregory, et al. 1993). EPA could then conclude that the value of the system in
question is X. However, EPA may be required to repeat this procedure at other sites to determine, in relative terms, how significant this value
(of X) is.

Status of the Method

Past studies and applications of this approach have focused primarily on group decision-making contexts where there is a need to
evaluate a range of management options and select the one that seems like it will perform the best across the attributes judged by decision
makers to be most important. The method has been applied in experimental studies in which people have been asked to evaluate its
effectiveness across a range of criteria that include the self-ratings of decision makers and measures of internal consistency (i.e., the degree to
which the approach helps people make choices that reflect their weighting of attributes) in choice (Arvai & Gregory 2003a, Arvai, et al. 2001).
The method has also been applied in a variety of practical contexts, including the setting of a national energy policy in Germany (Keeney, et al.

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1990), provincial water use planning in Canada (McDaniels, et al. 1999), and the management of a protected estuary (Gregory & Wellman
2001).

The goal of this discussion, however, is not to provide guidance about how EPA should make decisions; such advice falls outside the
charge of this committee. Instead, the goal is to highlight how these methods, which decompose complex decision problems and help people
carefully evaluate an option or range of options, may also be used for valuing the benefits of ecological systems and services. Because
decision-aiding methods are designed to help people to evaluate and then rank options, they may also be used to evaluate an environmental
system across a range of attributes and make judgments about its value relative to other systems, or indeed the same system at a previous point
in time. The method may also be combined with insights from multi-attribute utility theory to construct a single, uni-metric "value" that
encompasses the diverse array of attributes.

Strengths/Limitations

The strength of this method rests in its ability to not only integrate multiple attributes value, but also engage a broad spectrum of
stakeholders, holders of traditional ecological or cultural knowledge, and technical experts in the valuation process. In doing so, the method
has a high potential for identifying changes in ecosystems and their services that are likely to be of greatest concern to people. Moreover, by
engaging this broad spectrum of people, there is a greater likelihood that the valuation process will include attributes that wouldn't normally be
included by EPA, as well as those that may not easily be addressed by more traditional valuation approaches. Thus, this method may potentially
overcome (primarily) public or stakeholder objections to other approaches that are not perceived to adequately include moral and other non-
monetary aspects of value.

It is important to note, however, that the trade-offs, which are an important part of this process, are typically not easy to make. But,
because they are not holistic judgments that require the simultaneous integration of the various attributes, the likelihood that people will fail to
consider important attributes is low. Moreover, despite the effort that is required from those who use these methods, past experience suggests
that the outcomes are both more easily understood by people, and met with higher levels of support and ratings of defensibility when compared
with unstructured or unimetric approaches (Arvai 2003, Arvai & Gregory 2003b, Arvai, et al. 2001).

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As with many of the methods discussed in this report, this one requires that resources—time and expertise—be devoted to implementing it.
Engaging with stakeholders and technical experts to identify attributes that will be the focus of analysis, collecting data that characterizes these
attributes, and the process of making trade-offs all will require effort on the part of EPA.

Research Needs

As the primary focus of this method has been on providing decision support, its usefulness—particularly to potential users of the
method—as a complement to other valuation methods is unclear. For example, one wonders about its usefulness, in the context of many EPA
applications such as benefits assessment as mandated by OMB. Other questions can be raised about the effect of facilitation on the process as
one cannot guarantee that repeated applications of the process will produce the same outcomes. This question is not unique to decision aiding,
however, as a variety of factors (e.g., contextual, temporal, and spatial differences) may adversely affect other valuation methods as well.

References Providing Examples of Applications

Arvai, J., and R. Gregory. 2003a. A decision focused approach for identifying cleanup priorities at contaminated sites. Environmental Science
& Technology 37:1469-1476.

Arvai, J. L. 2003. Using risk communication to disclose the outcome of a participatory decision making process: Effects on the perceived

acceptability of risk-policy decisions. Risk Analysis 23:281-289.

Arvai, J. L., and R. Gregory. 2003b. Testing alternative decision approaches for identifying cleanup priorities at contaminated sites.

Environmental Science & Technology 37:1469-1476.

Arvai, J. L., R. Gregory, and T. McDaniels. 2001. Testing a structured decision approach: Value-focused thinking for deliberative risk

communication. Risk Analysis 21:1065-1076.

Gregory, R., J. L. Arvai, and T. McDaniels. 2001a. Value-focused thinking for environmental risk consultations. Research in Social Problems
and Public Policy 9:249-275.

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Gregory, R., T. McDaniels, and D. Fields. 2001b. Decision aiding, not dispute resolution: Creating insights through structured environmental
decisions. Journal of Policy Analysis and Management 20:415-432.

Gregory, R., and K. Wellman. 2001. Bringing stakeholder values into environmental policy choices: A community-based estuary case study.
Ecological Economics 39:37-52.

McDaniels, T., R. Gregory, and D. Fields. 1999. Democratizing risk management: Successful public involvement in local water management
decisions. Risk Analysis 19:497-510.

References

Arvai, J., and R. Gregory. 2003a. A decision focused approach for identifying cleanup priorities at contaminated sites. Environmental Science
& Technology 37:1469-1476.

Arvai, J. L. 2003. Using risk communication to disclose the outcome of a participatory decision making process: Effects on the perceived
acceptability of risk-policy decisions. Risk Analysis 23:281-289.

Arvai, J. L., and R. Gregory. 2003b. Testing alternative decision approaches for identifying cleanup priorities at contaminated sites.
Environmental Science & Technology 37:1469-1476.

Arvai, J. L., R. Gregory, and T. McDaniels. 2001. Testing a structured decision approach: Value-focused thinking for deliberative risk
communication. Risk Analysis 21:1065-1076.

Clemen, R. T. 1996. Making Hard Decisions: An Introduction to Decision Analysis. PWS-Kent Publishing Co., Boston, MA.

Gregory, R., J. L. Arvai, and T. McDaniels. 2001a. Value-focused thinking for environmental risk consultations. Research in Social Problems
and Public Policy 9:249-275.

Gregory, R., S. Lichtenstein, and P. Slovic. 1993. Valuing environmental resources: A constructive approach. Journal of Risk and Uncertainty
7:177-197.

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Gregory, R., T. McDaniels, and D. Fields. 2001b. Decision aiding, not dispute resolution: Creating insights through structured environmental

decisions. Journal of Policy Analysis and Management 20:415-432.

Gregory, R., and K. Wellman. 2001. Bringing stakeholder values into environmental policy choices: A community-based estuary case study.
Ecological Economics 39:37-52.

Keeney, R., D. von Winterfeldt, and T. Eppel. 1990. Eliciting public values for complex policy decisions. Management Science 36:1011-1030.
Keeney, R. L. 1992. Value-focused Thinking. A Path to Creative Decision Making. Harvard University Press, Cambridge, MA.

Keeney, R. L., and R. Gregory. 2005. Selecting attributes to measure the achievement of objectives. Operations Research 53:1-11.

Keeney, R. L., and H. Raiffa 1993. Decisions with multiple objectives: Preferences and value tradeoffs. Cambridge University Press,
Cambridge, UK.

McDaniels, T., R. Gregory, and D. Fields. 1999. Democratizing risk management: Successful public involvement in local water management
decisions. Risk Analysis 19:497-510.

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1	METHODS USING COST AS A PROXY FOR VALUE

2

3	Cost as a proxy for value, including replacement cost, tradable emissions permits, and habitat equivalency analysis (HEA), are a

4	distinct category of methods that use information about the cost of alternative means of providing the same quantity and quality of

5	ecosystem services to infer the value of protecting one particular means of providing the ecosystem services. However, because costs and

6	values are two distinct notions, great care needs to be taken in the application of these methods and in the interpretation of results using

7	these methods.

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Replacement Costs

Brief description of the method

This method, also called avoided cost, uses the cost of replacing ecosystem services with a human-engineered system as an estimate
of the value of providing ecosystem services via protection of an ecosystem. For example, an estimate of the value of conserving an
ecosystem that serves as a watershed that naturally provides clean drinking water could be derived by estimating the cost of building a water
filtration plant that would provide the same quantity and quality of water. Replacement cost is exactly what it says: the cost of replacing an
ecosystem service via some other means. Replacement cost is not a measure of the value of the ecosystem services themselves. Rather, it
is the value of having one particular means of providing ecosystem services, and therefore not having to pay to replace services via some
other means. Also, the replacement cost method should not be confused with applications of "averting behavior" based upon observed
voluntary behavior on individuals (see revealed preference methods).

Status as a method

The method has been used to provide estimates of the value of protecting watersheds for the purpose of providing clean drinking
water (NRC 2004). The most famous of such cases, and the example of valuing ecosystem services that is cited probably more than any
other, is the case of protecting the Catskills watersheds that provide drinking water for New York City (Chichilnisky and Heal 1998, NRC
2000, 2004). New York City, faced with the possibility of being required by EPA to build a water filtration plant for water from the
Catskills, opted to invest in greater watershed protection in the Catskills. New York City and EPA signed a Watershed Memorandum
Agreement in 1997 that allowed New York City to pursue a watershed protection plan in lieu of building filtration. While commonly cited
as a classic case of the value of protecting ecosystems, this case is not without controversy. It is not clear that protecting watersheds will
ultimately be successful in maintaining drinking water quality, or that the protection of watersheds versus building a filtration plant will
provide equivalent water quality in all dimensions (NRC 2004). Further, some analysts have suggested that the threat of building the
filtration plant had more to do with government regulations than with real water quality issues (Sagoff 2005).

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Another example using a replacement cost approach is the avoided cost of illness approach that EPA has used successfully to
account for certain human health benefits of environmental regulations.

Strengths/Limitations

Replacement cost can be a valid measure of value if three conditions are met: a) the human-engineered system provides services of
equivalent quality and magnitude; b) the human-engineered system is the least costly alternative; and c) individuals in aggregate would be
willing to incur these costs rather than forego the service (Bockstael, et al. 2000, Shabman and Batie 1978). If these conditions are not met,
then use of replacement cost is invalid. Even when these conditions are met, replacement cost, rather than being a value of ecosystem
services themselves, is the value of having a means to produce the service via an ecosystem instead of through an alternative human-
engineered system.

All valuation methods can be applied incorrectly and misinterpreted, but the replacement cost method requires special caution.
Because there is great potential for abuse in using replacement costs to estimate the value of ecosystem services, it should be used with care.
The loss of an ecosystem service does not necessarily mean that the public would be willing to pay for the least cost alternative. Similarly,
a regulatory constraint requiring replacement in the event of loss of ecosystem service also does not guarantee that the public would be
willing to pay to replace the service. If the value of the service does not exceed the cost of alternative means of providing the equivalent set
of services, then use of replacement cost is invalid. Even when the benefits of the service exceed the least cost method of providing the
service, replacement cost does not measure the willingness to pay for an environmental improvement or the avoidance of harm. It merely
represents the value (avoided cost) of not having to provide the service via human engineered approaches. Still, if there are alternative ways
of producing the same service, and if that service would be demanded if provided at the least cost human-engineered alternative method,
then replacement cost is a valid measure of the change in value from loss of the service provided by the ecosystem.

Key References

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1	Bockstael, N. E., A. M. Freeman, et al. (2000). "On measuring economic values for nature." Environmental Science and Technology 34:

2	1384-1389.

3	Chichilnisky, G. and G. Heal. (1998). "Economic returns from the biosphere." Nature 391: 629-630.

4	National Research Council (2000). Watershed Management for Potable Water Supply: Assessing the New York City Strategy.

5	Washington, D.C., The National Academies Press.

6	National Research Council (2004). Valuing Ecosystem Services: Toward Better Environmental Decision-Making. Washington, D.C., The

7	National Academies Press.

8	Sagoff, M. (June 2005) The Catskills parable. PERC Report. Bozeman, MT: Political Economy Research Center.

9	Shabman, L. A. and S. S. Batie (1978). "The Economic Value of Coastal Wetlands: A Critique." Coastal Zone Management Journal 4(3):
10 231-237.

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Tradable Permits

In the case of tradable permits, there are no conditions under which the cost of permits could be used as a proxy for economic value.

Emissions permit trading has been allowed under the Clean Air Act since the 1990 Amendments. Under a cap-and-trade system,
such as that used by EPA to reduce sulfur dioxide emissions, the regulatory body determines the total number of permits available and some
means of allocating permits among regulated sources. A regulated source must ensure that it has sufficient permits to cover its activities, or
it will face penalties. In the example of tradable emissions permits, a regulated source can take actions to reduce its own emissions and/or
purchase permits from other sources. For those firms with higher marginal cost of pollution control, cost savings can occur if they purchase
emissions-reduction credits from firms with lower pollution control costs. Similarly, firms with relatively low pollution control costs can
profit by undertaking greater abatement and selling extra permits. In so doing, trading can reduce overall costs of compliance. Tradable
permits schemes have been proposed in fisheries management in the form of individual transferable quotas (ITQs), and in land conservation
in the form of transferable development rights (TDRs).

It has been suggested that the price of a tradable permit is a proxy for the economic value of provision of environmental quality or
conservation. However, this confuses the notion of costs and benefits. In market equilibrium, the price of a tradable permit is equal to the
marginal cost of supplying a unit of environmental quality or conservation covered by the permit. Permit price need not bear any relation to
benefit of environmental quality or conservation. If there are a large number of permits issued relative to demand for permits then permit
price will be low; with few permits, price will be high. This does not necessarily mean that the value of environmental quality or
conservation is low (or high). Permit price only reflects value if price equals the marginal benefit of environmental improvement or
conservation, which occurs only if the number of permits issued is such that marginal costs and marginal benefits are equal. But issuing the
right number of permits to get marginal cost equal to marginal benefit requires knowing marginal benefit in the first place. There is no way
to be confident that tradable permit prices reflect value without already knowing value. In other words, tradable permit prices do not
constitute a valuation methodology capable of generating information about values.

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Habitat Equivalency Analysis

Brief description of the method

Habitat Equivalency Analysis (HEA) is an analytical framework originally developed to
calculate compensation for loss of ecological services resulting from injury to a natural resource
over a specific interval of time (King and Adler 1991, NOAA 1995). Figure 10 provides a
graphic representation of the relationship between the interim lost from an environmental
incident or activity and the recovery of the environment over time both due to natural
mechanisms and from primary restoration actions.

enhanced to replace an equivalent level of ecological services over time as were lost due to the
injury. The basic HEA formula is shown in Text Box 20. Ultimately the HEA approach is not a
valuation method but rather more appropriately defined as a "cost-replacement" method. Yet it
is important to recognize that an implicit operational assumption for an HEA is that the quantity
of ecological service flows, and their as yet undefined value, associated with any given unit of
lost or injured habitat are equivalent (same type and comparative value) to a unit of the proposed
replacement habitat.

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Text Box 20: Equation for Habitat Equivalency Analysis

'Debit: PDV Loss"	"Credit: PDV Gain"

-A	 	A_

r	>> r	^

X 1,(1+ /)<"-'>= X J!,(l + 0

t = t 0	S = S

where

o

Lt =	lost services at time t

Rv	replacement services at time .s-

to	t m when lost services are first

t;	time when lost services are last

so	time when replacement services are first

si	time when replacement services are last

P	nresent time when the natural resource damage claim is

i	periodic discount

There are two main steps in a HEA which are accomplished simultaneously: a)
quantifying the injury, and b) scaling the size of restoration to compensate for the lost service
over time due to that injury. To be clear, injury is not determined in a HEA, but such a
determination of injury is a necessary pre-step to provide the input for scaling the restoration to
match the degree of injury. The HEA approach focuses on scaling replacement costs on a
service-to-service basis. Therefore in quantitative expressions HEA relies on biophysical units
such as acres of habitat as a surrogate of service, and calculates the increase in habitat over time
in service acre years. A similar methodology, Resource Equivalency Analysis (REA) focuses on
scaling replacement costs on a resource-to-resource approach. In this context, resources are
generally defined in terms of biotic type and mass (e.g., kilograms of fish) for the quantification
of injury, but often ultimately revert back to an estimate of habitat required to replace or generate
those lost resources in estimating the size and type of replacement actions required to restore the
environment. HEA can also handle injuries to biotic resources but needs to equate those
resource losses to the unit of habitat it would take to create or support that mass of birds, fish,

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and invertebrates in the first place. Those performing an HEA will thus need to be careful in this
translation to avoid the potential for double counting if they are estimating habitat needs for
species which are supported by a common habitat such as coastal wetlands.

Temporal assumptions are very important in working with HEA, especially in a damage
assessment. Questions such as the following need to be answered or estimated:

•	How long has the injury or lost service been in place?

•	How much time is required to implement the restoration project?

•	How long will the restoration project take before it reaches full replacement service?

Obviously, the answers to these questions can have a significant impact on the estimated
compensatory value required to offset the injury. In HEA, a discount rate must be selected for
the Net Present Value calculations.

There are some crucial assumptions associated with the HEA method. It can be used
only when values per unit of replacement services and lost services are comparable, when it is
possible to use a common metric to define an injury and the value of replacement services, and
when replacement of ecological services is feasible and measurable.

Since HEA is a restoration/compensation method that is projected into the future, the
final unit is a Net Present Value (NPV) measure of the services in the future stated in discounted
terms (e.g., Discounted Service-Acre Years or DSAYs). Discounting or scaling of the
equivalency of any given sets of injured or restored habitat is required since the resource types
that are being addressed are not static over time (NOAA 1999). Injured resources can recover to
baseline conditions on their own and planted habitat takes time to develop to full maturity. So
factors such as baseline conditions and recovery times become key opportunities for uncertainty
in any HEA. Additionally for HEA to operate effectively it must fully explore and determine
that capacity of any project or suite of projects to achieve the required level of restoration. To
accomplish this assurance step, in advance of an HEA, a process referred to as C.O.P.E. was
developed (King 1997). The acronym C.O.P.E. stands for the attributes desired in the HEA,
which are: a) Capacity to provide service; b) Opportunity for project(s) in the correct location; c)
Payoff of comparable services; and d) Equity to provide service to people in the location that
suffered the injury. Each restoration project must satisfy the presumptions of C.O.P.E. to be

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worth further quantification via HEA as a contribution to satisfy the needed service years
equivalent to the lost interim service.

Example of how the method could be used as part of the C-VPESS expanded & integrated
framework

The spatial scale at which HEA has typically operated has been at the level of local to
regional decisions. Therefore it is not reasonable in its current state of development for HEA to
be considered as a tool useful for creating input to national rule-making. HEA also operates over
past and future time scales, involving compensation for injury or estimate service produced by
past action, as well as allowing time for restoration projects to mature to full ecosystem service
capacity.

With regard to where to place HEA in the C-VPESS integrated framework, it would seem
to bridge a number of the process elements. Although it would not be fair to say that it is
currently applied in a manner that would be classed as characterizing value, it does provide a
framing for characterizing bio-physical change. The HEA methodology relies on structural or
spatial measures of ecological components such as acres of habitat. Specific service categories
such as provisioning, regulating, cultural, and supporting services as expressed in the
Millennium Ecosystem Assessment framework (2005) are not identified or expressed but would
be considered to be present and operating But if the type of habitat or resources can, with further
research, be equated to a unitized measure of values or service flows, either monetary or
otherwise, then HEA could be used to scale that associated value over time and across alternative
actions. If, through research and development, service flows and associated values can be
quantified for given habitat categories (e.g., an acre of coastal wetlands in Louisiana), then there
is some hope that HEA may evolve to be a support for valuation.

Additionally, although HEA and REA are currently used in the post-hoc context of
injury, damages, and compensation, there is no reason that these methods are constrained to
managing adverse outcomes after the fact. They could just as easily be used ex ante to compare
alternative future actions to identify the action with the least impact and to compare alternative
actions to identify which will yield the most service or equal service in the shortest time frame.
These methods or variations could be a fruitful avenue for the Agency to explore through its
research and development activities.

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As noted, HEA is a tool that has application constraints. Typically, the HEA is applied to
support local decisions by scientific experts to evaluate project alternatives for achieving
restoration objectives. Such analyses allow those experts to arrive at convincing trades among
restoration options. Although there is not much evidence to indicate the use of HEA in support
of a facilitated or mediated process that includes the general public, there do not appear to be any
technical reasons why this could not be a useful application of HEA to project the services
provided by possible alternative future scenarios resulting from a suite of restorations actions.
Such engagement of the public in the identification of restoration projects and desired services is
likely to leadto more widely accepted restoration decisions.

Status as a method

The HEA approach was originally developed in 1992 to quantify damages associated
with contaminated wetlands (King and Adler 1991, Malcolm v. National Gypsum 1993 as
referenced in Unsworth and Bishop 1993) and has since been applied to cover injuries due to
chronic contamination, spills, and vessel groundings in a variety of habitats (Chapman, et al.
1998, Fonseca, et al. 2000, Milon and Dodge 2001, NOAA 2001). HEA is currently used in
Natural Resource Damages Assessment (NRDA) under Oil Pollution Action (OPA) and
CERCLA (Superfund). The purpose of NRD actions is to make the public's interest whole for
injuries to natural resources that result from the release of hazardous substances or oil. It is
important to note that restoration for damages is distinct from remediation activities.

Interestingly, under these two regulatory frameworks there is a different focus on
compensation. Under Superfund actions, compensation for damages is focused on monetary
compensation, which requires restoration of service ultimately to be converted to replacement
costs in dollars. Under OP A, the focus is on replacement of resources to achieve compensation.
The question is how much in the way of new public resources does the public require to be made
whole for their loss. Therefore, value is scaled from resource or habitat lost to resource or habitat
replaced. As noted previously, there are no barriers to applying these methods in proactive
support of decisions. Therefore the Agency should explore such proactive applications of HEA
and REA in other regulatory contexts and especially in collaborative partnerships with
conservation as a focus.

Strengths/Limitations

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The HEA method can be used as a way to scale surrogates measures (e.g. acres of
Habitat or mass of fish) of non-market services often overlooked by other valuation methods
when the specific assumptions associated with HEA can be met. The method is not complicated
mathematically. It is by nature inter-disciplinary because determination of comparability per
unit of replacement services and lost services requires collaboration between ecologists and
economists.

Since HEA and REA are currently applied to support regulatory actions which link to a
litigation process, to define compensation the analysis and supporting data need to be legally
defensible with regards to analytical quality. The chief analytical difficulty is to determine
defensible input parameters, especially an appropriate metric for lost and restored services and
related time functions for recovery and development to maturity.

The HEA method is not appropriate for standard benefit-cost analysis, where the goal is to
determine optimal (efficient) allocation of scarce resources. The cost of compensatory
restoration projects should not be communicated as the benefit of the resources to the public.
Treatment of Uncertainty

Uncertainty can be, and should be, directly incorporated into any HEA analysis.
Addressing uncertainty in inputs (e.g., percent service lost per unit of habitat and recovery time)
can be effectively done. Tracking the effects of uncertainty on HEA outputs can be easily
performed. One of the benefits of HEA is the transparency of the method. Sensitivity and
uncertainty analysis can be directly incorporated into a HEA evaluation and the resulting change
can be tracked in outputs (see NOAA 1999 for more details)

Research needs

There are a number of key areas for research and development that the Agency should
explore in connection with HEA.

The Agency should look at HEA for its applications in contexts other than Natural
Resource Damage Assessment. In particular, it should consider its utility tandem with Net
Environmental Benefit Analysis (Efroymson, et al. 2004) in the selection of best alternatives for
project investment.

The Agency should consider research to develop a more complete understanding of the
service flows and the associated values of goods and services derived from those flows in
specific important habitat types (e.g., coastal wetlands, bottomland hardwood forests). Such

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value definitions for ecosystem service could then be coupled to HEA to estimate values

associated with a project or restoration action.

EPA should consider developing operating principles for considering on-site, in-kind

changes in resources and ecological services, as compared with off-site and out-of-kind

resources. In support of this objective, methods to assess and compare ecological capacity and

the opportunity and payoff for restoration in the evaluation and design of restoration projects will

also strengthen the method to assess comparability of ecological resources.

Finally, this method will be strengthened if the Agency develops guidance on the

appropriate aggregation and accounting of services related to biotic resources and their

supporting habitats in order to advance the utility of HEA to support local and regional valuation

efforts.

Key References

US ACE - Wetlands Permitting Reference

Chapman, D., N. Iadanza, and T. Penn. 1998. Calculating Resource Compensation: An

Application of Service-to-Service Approach to the Blackbird Mine Hazardous Waste
Site. National Oceanic and Atmospheric Administration, Damages Assessment Center .
Technical Paper Series 97-1, October 16, 1998 17 pp.

Dunford, R.W., T.C. Ginn and W.H. Desvousges. 2004. The Use of Habitat Equivalency

Analysis in Natural Resource Damage Assessment. Ecological Economic. Volume 48: pp
49-70

Fonseca, M.S., B.E. Julius and W.J. Kenworthy. 2000. Integrating Biology and Economics in

Seagrass Restoration: How Much is Enough and Why? Ecological Engineering. Volume
15. Pages 227-237

King, D.M. 1997. Comparing Ecosystem Services and Values: With Illustrations for Performing
Habitat Equivalency Analysis. National Oceanic and Atmospheric Administration. Silver
Spring, MD. Service Paper Number 1

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King, D.M. and K.J. Adler. 1991. Scientifically Defensible Compensation Ratios for Wetlands
Mitigation. U.S. Environmental Protection Agency Office of Public Policy, Planning and
Evaluation, Washington, DC. 16 pp., 3 figures and 2 tables, January 1991

Malcolm v. National Gypsum Co. 995 F.2d 346, 352 (2d Cir 1993)

Milon, J.W. and R.E. Dodge. 2001. Applying Habitat Equivalency Analysis for Coral Reef

Damage Assessment and Restoration. Bulletin of Marine Science. Volume 69, Number 2.
Pages 975 - 988

NOAA. 1995. Habitat Equivalency Analysis: An Overview. Policy and Technical Paper Series
No. 95-1 (revised 2000). National Oceanic and Atmospheric Administration,
Washington, DC

NOAA, 1999. Discounting and the treatment of Uncertainty in Natural Resource Damage

Assessment. Technical Paper 99-1 National Oceanic and Atmospheric Administration,
Damage Assessment and Restoration Program, Damage Assessment Center, Resource
Valuation Branch, Silver spring Maryland.

NOAA, 2001, Damage Assessment and Restoration Plan and Environmental Assessment for the
Point Comfort/Lavaca Bay NPL Site Recreational Fishing Service Losses, National
Oceanic and Atmospheric Administration et al.. 2001. Washington, DC

Millenium Ecosystem Assessment. 2005 . Ecosystems and Human Well-Being: Synthesis. Island
Press. Washington, D.C. 137pp.

Unsworth and Bishop 1993

Internet

NOAA Coastal Service Center - Habitat Equivalency Analysis -
www.csa.noaa.gov/economics/habitatequ.htm

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APPENDIX C: SURVEY ISSUES FOR ECOLOGICAL VALUATION:
CURRENT BEST PRACTICES AND RECOMMENDATIONS FOR

RESEARCH

Survey methods support many of the approaches for eliciting and measuring information
about values discussed in the C-VPESS report. Although scientific and technical issues
concerning survey design and administration can affect some aspects of ecological valuation,
they are distinct from the science and value assessment issues that are the main focus of the C-
VPESS report.

The C-VPESS recognizes, however, that issues related to survey methods are important
to some methods of ecological valuation and learned they were of particular concern to EPA
representatives participating in the SAB's December 13-15, 2005 Workshop, "Science for
Valuation of EPA's Ecological Protection Decisions and Programs." After that workshop, the
committee requested that this appendix be commissioned to supplement the main body of the
committee's report. This appendix provides an introduction for EPA Staff to questions posed to
the C-VPESS pertaining to survey use for ecological valuation. It provides an overview of how
recent research and evolving practice relating to those questions might assist the Agency.

Defining Survey Research

Survey research entails collecting data via a questionnaire from a sample of elements
(e.g., individuals or households) systematically drawn from a defined population (see Babbie,
1990; Fowler, 1988; Frey, 1989; Lavrakas, 1993; Weisberg, et al., 1996).46 Conducting a survey
involves (1) drawing a sample from a population, (2) collecting data from the elements in that
sample, and (3) analyzing the data generated. Survey research is a well-established and
respected scientific approach to measuring the behavior, attitudes, and beliefs, and much more of
populations of individuals.47 Surveys are usually done for one of three reasons: (1) to document
the prevalence of some characteristic in a population, (2) to compare the prevalence of some
characteristic across subgroups in a population, and/or (3) to document causal processes that
produce behaviors, beliefs, or attitudes. Because scientific surveys involve probability sampling,
their results can be used to estimate population parameters. This appendix addresses issues of

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survey methodology that cut across many different applications, including monetary valuations
(e.g., CVM), measures of preference, importance or acceptability, and determinations of the
assumptions, beliefs and motives that might underlie these expression of value.

Designs of Surveys

Surveys can take on a variety of designs, which are suitable for addressing different types
of research questions. For example, cross-sectional surveys are useful for measuring a variable
at a given point in time, whereas repeated cross-section surveys are more useful for observing
change over time in a population, panel surveys are more useful for examining change over time
in a sample of respondents, and surveys that implement experiments may be more useful for
establishing causality, although many types of information can be derived from the data from
each of these types of surveys.

Cross-sectional surveys involve the collection of data at a single point in time from a
sample drawn systematically from a population and are often used to document the prevalence of
particular characteristics in a population. Cross-sectional surveys allow researchers to assess
relations between variables and differences between subgroups of respondents. Data from cross-
sectional surveys can also be used to provide evidence about causal hypotheses using statistical
techniques (e.g., two-stage least squares regression or path analysis; Baron & Kenny, 1986;

James & Singh, 1978; Kenny, 1979), by identifying moderators of relations between variables
(e.g., Krosnick, 1988), or by studying the impact of an event occurring in the middle of data
collection (e.g., Krosnick & Kinder 1990).

Repeated cross-sectional surveys involve collecting data from independent samples
drawn from the same population at two or more points in time. Such data can be used to provide
evidence about causality, by gauging whether changes in an outcome variable parallel changes in
a purported cause of it. Repeated cross-sectional surveys can also be used to study the impact of
social events that occurred between the surveys (e.g., Weisberg, et al., 1995).

Panel surveys involve collecting data from the same sample of respondents at two or
more points in time and can be used to gauge the stability of a construct over time and identify
the determinants of stability (e.g., Krosnick, 1988; Krosnick & Alwin, 1989). Panel surveys can
also be used to test causal hypotheses, by examining whether changes over time in a purported
case correspond to changes in an outcome variable, by assessing whether changes over time in

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the outcome variable can be predicted by prior levels of the purported cause, or by testing the
effects of events that occur between waves (see, e.g., Blalock, 1985; Kessler & Greenberg, 1981,
on the methods; see Rahn, et al., 1994, for an example).

Panel surveys also face a number of challenges. Respondent attrition (or "panel
mortality") occurs when some of the people who provide data during the first wave of
interviewing are unreachable or refuse to participate in subsequent waves. Attrition reduces a
panel's effective sample size and it is particularly undesirable if a non-random subset of
respondents drop out. However, the literature on panel attrition suggests that panel attrition
minimally affects sample composition (Becketti, et al., 1988; Clinton, 2001; Falaris & Peters,
1998; Fitzgerald, et al., 1998a; 1998b; Price & Zaller, 1993; Rahn, et al., 1994; Traugott, 1990;
Zabel, 1998; Zagorsky & Rhoton, 1999; and Ziliak & Kniesner, 1998 ; although see Groves, et
al., 2000; Lubin, et al., 1962; and Sobel, 1959).

A second methodological issue in panel research is panel conditioning, or the possibility
that interviewing people repeatedly may change them and thereby make the sample less
representative of the larger population to which investigators wish to generalize. But again, the
literature on these issues is reassuring for the most part. A number of studies have found either
no evidence of panel conditioning effects or very small effects (Clinton, 2001; Cordell &

Rahmel, 1962; Himmelfarb & Norris, 1987; Sobol,1959; Willson & Putnam, 1982). Particularly
if repeated interviews with panel members touch on a wide variety of topics, each wave may
blend in with memories of prior waves via what psychologists call "retroactive interference,"
thus minimizing the likelihood of stimulated interest in any one topic. However, some evidence
suggests that interviewing people on a particular topic may cause them to become more
cognitively engaged in that topic (Bridge, et al., 1977; Granberg & Holmberg, 1992; Kraut &
McConahay, 1973; Willson & Putnam, 1982; Yalch, 1976; although see Mann, 2005). Other
studies have documented that asking people just one question about their behavioral intentions
can affect their subsequent behavior (see, e.g., Greenwald et al., 1987; Gregory, et al., 1982).

Interestingly, membership in a long-term panel survey may actually be beneficial to the
quality of data collected because of "practice effects" (e.g., Chang & Krosnick, 2001). The more
a person performs any task, the more facile and effective he or she becomes at doing so. In our
case, the tasks of interest include question interpretation, introspection, recollection, information
integration, and verbal reporting (see Tourangeau, et al., 2000).

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Mixed designs are used when researchers can capitalize on the strengths of more than one
of these designs by incorporating elements of two or more into a single investigation. If, for
example, a researcher is interested in conducting a 2-wave panel survey but is concerned about
conditioning effects, she could also administer the wave 2 panel questionnaire to an independent
cross-sectional sample drawn from the same population at the time of the second wave.
Differences between the data collected from the two wave 2 samples would suggest that carry-
over effects were, in fact, a problem in the panel survey.

Experiments can also be implemented in surveys to test causal hypotheses. If
respondents are randomly assigned to "treatment" and "control" groups that are asked different
versions of a question or question sequence, differences between the two groups can then be
attributed to the treatment.

Elements of a Weil-Defined Survey
Sampling

When designing a survey's sample, the sampling frame (the complete list of elements in
the population to which one wishes to generalize findings) must be defined, and the subset of
elements (the individual unit about which information is sought) in the population to be
interviewed must be selected. These decisions have important implications for the results of the
survey because they may impact both coverage and sampling error (see, e.g., Laumann, et al.,
1994). Coverage error occurs when the sampling frame excludes some portion of the population.
For example, telephone surveys usually exclude households without telephones. Sampling error
is the discrepancy between the sample data and the true population values that is due to random
differences between the sample and the sampling frame.

There are two broad classes of sampling methods: nonprobability and probability
sampling. Nonprobability sampling refers to selection procedures such as haphazard sampling,
purposive sampling, snowball sampling, and quota sampling in which elements are not randomly
selected from the population or in which some elements have zero or unknown probabilities of
selection. Probability sampling refers to selection procedures such as simple random sampling,
systematic sampling, stratified sampling, or cluster sampling in which elements are randomly
selected from the sampling frame and each element has an independent, known, nonzero chance
of being selected. Unlike nonprobability sampling, probability sampling allows researchers to be

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confident that a selected sample is representative of the population from which it was drawn and
to generalize beyond the specific elements included in the sample. Probability sampling also
allows researchers to estimate sampling error, or the magnitude of uncertainty regarding obtained
parameter estimates. Therefore, the best survey designs (and virtually all scientific surveys) use
some form of probability sampling.

Sampling error can be minimized by surveying large samples. However, the relation
between sampling error and sample size is not linear. A moderate sample size reduces sampling
error substantially in comparison with a small sample size, but further increases in sample size
produce smaller and smaller decrements in sampling error. Thus, researchers should recognize
that beyond a moderate sample size, the funds necessary to produce a large sample might be
better spent reducing other types of error.

Questionnaire Design

Introduction. High-quality, scientific surveys typically provide respondents with several
key pieces of information when introducing the survey, whether it is through an introductory
mailed letter, an e-mail, or an introduction from a telephone or face-to-face interviewer. This
information protects respondents' rights, helping to ensure that the survey is being conducted
ethically, and it may help to increase the perceived validity of the survey and, as a result,
respondent participation. This information includes information about the sponsor of the survey,
a brief description of the topic of the survey, and how the data from the survey will be used. The
introduction should also include a reassurance to respondents that their survey responses will be
kept confidential and a description of any other measures in place to protect respondents.

Finally, the burden being placed on respondents and any risks to the respondent should also be
described. This information allows respondents to provide informed consent. That is, knowing
this information, respondents can make an informed choice about whether or not to participate in
the survey. However, it is important to also keep this introduction as short as possible, as longer
introductions place a greater burden on respondents and may also reduce survey participation.

Survey questions. All surveys include questions, and a series of decisions must be made
to achieve optimal designs of those questions. First, a researcher must decide if each question
will be open- or closed-ended. For closed-ended questions, a researcher interested in obtaining
rank orders of objects must decide whether to ask respondents to report those rank orders directly

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or to rate each object separately. If respondents are asked to rate objects, the researcher must
decide how many points to put on the rating scale, how to label the scale points, the order in
which response options will be offered, and whether respondents should be explicitly offered the
option to say they "don't know" or have no opinion. Once the questions are written, the
researcher must determine the order in which they will be administered. Researchers must also
decide how to optimize measurement on sensitive topics, where social desirability response bias
may lead respondents to intentionally misreport answers in order to appear more respectable or
admirable. A large body of relevant scientific studies about the questionnaire design decisions
faced by researchers has now accumulated, and when taken together, their findings clearly
suggest how to design questionnaires to maximize the quality of measurement. Although a
description of the entire literature is beyond the scope of this review, we provide a few examples
here about survey questions using rating scales to provide a flavor of what this literature has to
offer.

When designing a rating scale, one must begin by specifying the number of points on the
scale (for a review of relevant literature, see Krosnick & Fabrigar, forthcoming). For bipolar
scales, which have a neutral point in the middle (e.g., running from positive to negative),
reliability and validity are highest for about seven points (e.g., Matell & Jacoby, 1971). In
contrast, the reliability and validity of unipolar scales, with a zero point at one end (e.g., from no
importance to very high importance), seem to be optimized for somewhat shorter scales,
approximately 5 points long (e.g., Wikman & Warneryd, 1990).48

A number of studies show that data quality is better when all points on a rating scale are
labeled with words than when only some are labeled thusly (e.g., Krosnick & Berent, 1993).
Researchers should try to select labels that have meanings that divide up the continuum into
approximately equal units (e.g., Klockars & Yamagishi, 1988). For example, "very good, good,
or poor" is a poor choice, because the meaning of "good" is much closer to the meaning of "very
good" than it is to the meaning of "poor" (Myers & Warner, 1968).49

Researchers must then decide how to order the response alternatives, and people's
answers to rating scale questions are sometimes influenced by this order. After reading the stem
of most rating scale questions, respondents are likely to begin to formulate a judgment. For
example, the question, "How effective do you think the clean-up plan will be?" would induce
respondents to begin to generate an assessment of effectiveness. As respondents read or listen to

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the answer choices presented, some may settle for the first acceptable response option they
encounter rather than considering all the response options and selecting the answer choice that
best reflects their judgment, thus resulting in primacy effects in ratings, which have been
observed in many studies (e.g., Belson, 1966; Carp, 1974; Chan, 1991; Matthews, 1929). To
minimize bias, it is therefore usually best to rotate the order of response choices across
respondents and to statistically control for that rotation when analyzing the data.50

Pretesting. Even the most carefully designed questionnaires sometimes include items that
respondents find ambiguous or difficult to comprehend, or items that respondents understand, but
interpret differently than the researcher intended. Researchers can conduct pretests of a draft
questionnaire to identify these kinds of problems. Pretesting methods include conventional
pretesting, in which interviewers conduct a series of interviews and report any problems with
question interpretation or comprehension (see, e.g., Bischoping, 1989; Nelson, 1985); behavior
coding, in which a researcher notes the occurrence of verbal events during the interview that
might indicate problems with a question (e.g., Cannell, et al., 1981); and cognitive interviewing,
in which a questionnaire is administered to individuals who either "think aloud" while answering
or answer questions about the process by which they formulated their responses (e.g., Forsyth &
Lessler, 1991). Each of these methods has advantages and disadvantages. When resources are
available, researchers can use multiple methods to pretest questionnaires because different
methods identify different types of problems (see Presser et al., 2004).

Mode of Data Collection

Survey data can be collected in one of four primary modes: mail, telephone, face-to-face,
and Internet. Interviewers administer telephone and face-to-face surveys, whereas mail and
Internet surveys involve self-administered questionnaires. Mode choice can produce notable
differences in survey findings. So mode choice must be made carefully in light of each project's
goals, budget, and schedule. Each survey mode has advantages and disadvantages. When
choosing a mode for a particular survey, researchers must consider cost, characteristics of the
population, sampling strategy, desired response rate, question format, question content,
questionnaire length, length of the data-collection period, availability of facilities, the purpose of
the research, and the resources available to implement it.

Aspects of the population, including literacy, telephone coverage, and familiarity and

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access to computers, are important in the decision about mode. Literacy is necessary for self-
administered questionnaires. Broad telephone coverage of the population is necessary when
conducting a telephone survey. Internet access and familiarity with computers is important for
an Internet survey.

Coverage error is minimized in face-to-face household surveys, but is larger in Random
Digit Dial (RDD) telephone household surveys, because they exclude respondents without
telephones and those with only cell phones. Coverage error for mail and Internet surveys
depends upon the sampling strategy used and with list samples, the quality of the list that is used
as the initial sample frame.

Although probability sampling is possible in all modes, mode affects the ease with which
it can be implemented. Telephone and face-to-face surveys routinely use probability household
sampling strategies, but mail and other self-administered surveys are more commonly used when
a list of the entire population is available. In some Internet surveys, nonprobability sampling
methods are used (e.g., inviting individuals to opt in through websites), which does not yield
results that can be generalized to the population of interest (Malhotra & Krosnick, in press).

Some researchers, however, have implemented probability sampling to recruit respondents to
complete questionnaires weekly via the Internet and provided Internet access to respondents who
do not have it.

Mode also influences the response rates achieved in a survey, with face-to-face surveys
typically achieving the highest response rates. Telephone surveys achieve somewhat lower
response rates, and self-administered mail surveys achieve low response rates unless a sequence
of multiple contacts are implemented at considerable cost and with considerable implementation
time (see Dillman, 2006).

The types of information and questions researchers wish to present may also influence
the choice of mode. If a survey includes open-ended questions, face-to-face or telephone
interviewing is preferable because interviewers can probe incomplete or ambiguous respondent
answers. If complex information will be presented as part of the survey, face-to-face
interviewing or Internet questionnaires allow the presentation of both oral and visual
information. If the researcher needs to ask questions about sensitive topics, self-administered
questionnaires and computers provide respondents with a greater sense of privacy and therefore

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elicit more candid responses than interviewer-administered surveys (e.g., Bishop & Fisher, 1995;
Cheng, 1988; Wiseman, 1972). Face-to-face interviewing is likely to elicit more honest answers
than telephone interviewing because face-to-face interviewers can develop better rapport with
respondents and more easily implement private response methods.

Face-to-face data collection permits interviews of an hour or more, whereas telephone
interviews usually last no more than 30 minutes. With self-administered questionnaires,
response rates typically decline as questionnaire length increases, so they are generally kept even
shorter.

Telephone and Internet surveys can be completed in very short field periods, often within
a matter of days (though at the cost of lower response rates). In contrast, mail surveys require
significant amounts of time, and follow-up mailings to increase response rates further increase
the overall turnaround time. Similarly, face-to-face interview surveys typically require a
substantial length of time in the field.

Face-to-face interviews are usually considerably more expensive than telephone
interviews, which are usually about as expensive as self-administered questionnaire surveys of
comparable size using methods necessary to achieve high response rates. The cost of Internet
data collection from a probability sample is about equivalent to that of telephone RDD
interviewing.

These differences between modes also contribute to differences in data quality. Face-to-
face surveys have the highest response rates, are the most flexible in terms of interview length
and presentation of complex information, and acquire more accurate reports than do telephone
surveys (Holbrook, et al., 2003). Internet surveys allow presentation of complex information,
and reporting accuracy appears to be higher in Internet surveys than in telephone surveys (Chang
& Krosnick, 2001). Although response rates from Internet surveys based on initial RDD
telephone samples are quite low and have similar coverage error to telephone surveys, such
difficulties may be reduced by recruiting probability samples of respondents face-to-face in their
homes.

Assessing Survey Accuracy

In order to optimize survey design or to evaluate the quality of data from a particular

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survey, it is necessary to assess accuracy (or conversely error) in survey data. If optimal
procedures are implemented a high level of accuracy can be achieved, but departures from such
procedures can compromise the accuracy of a survey's findings. Usually, researchers have a
fixed budget and must decide how to allocate those funds in order to maximize the quality of
their data. According to the "total survey error" approach, a research can think about survey
design issues within a cost-benefit framework geared toward helping researchers make design
decisions that maximize data quality within budget constraints (cf. Dillman, 1978; Fowler, 1988;
Groves, 1989; Hansen & Madow, 1953; Lavrakas, 1993).

The total survey error perspective recognizes that the goal of survey research is to
accurately measure particular constructs in a sample of people who represent the population of
interest. In any given survey, the overall deviation from the ideal is the cumulative result of
several sources of survey error. The total survey error perspective disaggregates overall error
into four components: coverage error, sampling error, nonresponse error and measurement error.
Coverage and sampling error have previously been described. Nonresponse error is the bias
that can result when data are not collected from all members of a sample. Measurement error
refers to all distortions in the assessment of the construct of interest, including systematic biases
and random variance that can be brought about by respondents' own behavior (e.g., misreporting
true attitudes), interviewer behavior (e.g., misrecording responses), and the questionnaire (e.g.,
ambiguous or confusing question wording).

Nonresponse occurs when data are not collected from all of the eligible sample elements.
Nonresponse occurs either because sampled elements are not contacted (e.g., no one is ever
home) or because members of sampled households decline to participate. The response rate for a
survey is the proportion of eligible sample elements from whom data were collected and is
almost always less than 100%. Lower response rate increase the risk that the sample is not
representative of the population.

To maximize response rates researchers implement various procedures. For example, the
field period during which potential respondents are contacted can be lengthened (e.g., Groves &
Lyberg 1988; Keeter et al. 2000), the number of times an interviewer tries to contact a
household member can be increased (Merkle, et al., 1993; O'Neil, 1979), financial incentives can
be offered for participation (e.g., Singer et al., 1999; Singer, et al., 2000), advance letters can be

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mailed to households to inform residents about the survey (e.g., Camburn et al., 1995; Link &
Mokdad 2005), and the questionnaire can be kept as short as possible (e.g., Collins et al. 1988).
All of these strategies have been found to increase response rates in at least some studies in
which these factors were considered one-by-one. However, some strategies, such as sending
advance letters or leaving messages on potential respondents' answering machines, may not
always be successful because they give advance notice that interviewers will try to contact
respondents, and respondents may use this knowledge to avoid being interviewed.

Low response rates increase only the potential for nonresponse error, because
nonresponse error is a function of two variables: the response rate and the size of the difference
between respondents and nonrespondents. If respondents and nonrespondents do not differ
substantially, response rates will be unrelated to nonresponse bias. That is, it is possible to
conduct a survey with a response rate of 20% and end up with data that describe the population
quite accurately.

A number of publications using a variety of methods have shown that as long as a
representative sample is scientifically drawn from the population and professional efforts are
made to collect data from all potential respondents, variation in response rates (between 20% and
65%) does not substantially increase the accuracy of the survey's results (Curtin, et al., 2000;
Holbrook, et al., in press; Keeter, et al., 2000). Furthermore, although many surveys manifest
substantial non-response error, there is little evidence that the observed amount of nonresponse
error is related to the response rate for the survey.

Measurement error includes any distortion or discrepancy between the theoretical
construct of interest and the concrete measurement of that construct. One method for assessing
measurement error is to compare responses to a survey to a known standard to assess their
validity. For example, reports of whether or not a respondent voted in an election can be
compared to public records of voting, or reports of drug use can be compared to the results of
drug tests performed on hair, urine, or saliva samples. However, surveys often measure
constructs for which there are no available standards. In these cases, the reliability or predictive
validity of survey measures is often used to judge the quality of the measurement. One method
for comparing different survey questions or question orders is to use split-ballot experiments in
which half of respondents are randomly assigned to receive one form of a questionnaire (using

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one question wording or order) and the other half are randomly assigned to receive a different
form of the questionnaire (using a second question wording or order). One or more of the
approaches described above (e.g., comparison to a known standard, reliability or predictive
validity) can then be used to compare the reliability and/or validity of responses across
questionnaire form to determine if one question wording or order is better.

The total survey error perspective advocates explicitly taking into consideration each of
these four sources of error and making decisions about the allocation of resources with the goal
of reducing the total error. Many steps that do not cost real dollars can be taken to reduce error,
but other steps to reduce error do cost money, and the more money a researcher spends to reduce
one type of error, the less money he or she has available to reduce other types of error.
Researchers should make such tradeoffs explicitly, recognizing the opportunity costs they pay
when making a particular move to maximize quality in a particular way, selecting approaches
likely to yield the biggest bang for the buck spent.

Challenges in Using Surveys For Ecosystem Protection Valuation

Introduction. One application of the survey method is for assessing the value of
ecosystems and services. A variety of techniques have been developed to assess the monetary
value of ecosystems, and these values can be used as input to required cost-benefit analyses by
EPA in the policy-making process. When monetary values are not required or are too difficult to
attain or are deemed ethically or otherwise inappropriate to the problem at hand, surveys can be
used effectively to determine quantitative measures of preference, importance or acceptance of
alternative policies, actions and outcomes. When surveys are used for valuation, many
respondents are asked to rank, rate or place a monetary value on a change in ecosystems/services
conditions with which they may not be familiar prior to the survey, but this does not mean that
respondents lack a value for the ecosystem in question. Respondents' experiences have
cumulated into beliefs and attitudes stored in long term memory that are the ingredients of their
orientations toward objects they will encounter in the future. Therefore, an important component
of valuation survey design is to describe the ecosystem as fully as possible so that respondents
can use these beliefs and attitudes to determine its value. Doing so helps to maximize the extent
to which the values that respondents report validly reflect these underlying beliefs and opinions.
This means that valuation surveys will be different from most other surveys because they must

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devote a considerable amount of time to educating the respondent about the ecosystem in
question. This may require respondents to listen to or read relatively long passages of text and
perhaps to observe visual presentations of nonverbal information as well, such as charts, maps,
drawings, or photographs.

Conveying a large amount of information. It is important that the survey provides all of
the information that respondents want in order to make the judgments being asked of them and
present that information in a way that is understandable to all respondents. To achieve these
goals, researchers can begin by conducting research with pretest respondents to assess what
information they want to know and their understanding and interpretation of information
presented to them. These procedures can be used iteratively to refine the presentation to enhance
understanding and sufficiency of the information set.

In order to present a sizable set of information to respondents, a variety of techniques can
be implemented to maximize comprehension. The principles of optimal design can be used to
construct graphical displays of information (e.g., Kosslyn, 1994; Tufte, 2001). A great deal of
information can also be presented to respondents in a single visual display that a respondent can
read or an interviewer can explain to the respondent. Information can also be presented in the
narrative form of a story, for example, by telling respondents that they'll be told about:a) the
state of an ecosystem as it used to exist 50 years ago; b) changes that have occurred to the
ecosystem in the intervening years; c) the causes of those changes; d) what could be done to
reverse those changes; and e) how this could be implemented. Rather than lecturing respondents
for a long time period, a questionnaire can maintain respondent engagement by presenting
information in small chunks, separated by questions allowing respondents to react briefly to the
information they've been given (e.g., "Had you ever heard of the Golden River before today?").
Respondents can also be asked periodically to verbalize any information that they'd like to have
as the story progresses, to allow them to express their cognitive responses to the presentation.

The choice of survey mode also impacts the presentation of information about an
ecosystem. Face-to-face interviewing is optimal because it allows visual displays of any type
and interviewers can create a strong sense of interpersonal connection with respondents.
Telephone interviewing permits a similar connection, though probably less strongly, and visual
displays are usually not possible. Computer administration of a questionnaire can include static
and dynamic presentation of visual and aural information, and questions can be interspersed with

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this information, but it may not be possible to create the strong sense of connection between the
respondent and the researcher. Self-administered paper and pencil questionnaires allow only
visual presentation of information and do not allow information to be presented in small chunks
(because respondents can look ahead in the questionnaire). A large volume of information
presented densely on a large set of pages of paper may be intimidating or dispiriting, thus,
minimizing respondent motivation and provoking superficial processing of the information. The
self-administered mode may be the least desirable for this reason. For all modes, it is important
to pretest the final instrument to be sure it's working as intended.

Communicating uncertainty. Because of the uncertainty inherent in estimating the effect
a policy might have on an ecosystem or service (see Section 8.1), researchers using surveys for
valuation may not only want to convey large amounts of information to respondents, but they
may also want to convey their level of certainty or uncertainty about that information. Such
uncertainty could be conveyed to respondents in a number of ways, including providing ranges
or confidence intervals for the information provided (e.g., the estimated cost of maintaining the
ecosystem is between 1 and 3.3 million dollars per year), providing a verbal description of
scientists confidence in the information (e.g., scientists are very confident that a policy will
protect an ecosystem), communicating the degree of consensus about the information among
scientists (e.g., 75% of scientists agree that a particular policy will protect the ecosystem), or
conveying the probability that an outcome or benefit will occur (e.g., scientists believe this
policy has a 75% probability of protecting the ecosystem). There is substantial evidence that
people have difficulty the last type of evidence accurately (e.g., Tversky & Kahneman, 1974),
but the EPA may want to explore these various methods for conveying uncertainty to determine
the extent to which people understand and use different types of information about uncertainty in
valuation.

Scale and spatial issues. Because the spatial and temporal scale of ecological systems
and services may impact valuation processes, these dimensions should be incorporated into the
communication of information and the measurement of value. For example, the information that
respondents receive during the survey interview should, if possible, explicitly describe the scale
of a proposed policy or the ecosystem or service for valuation. This is particularly true if the
scale is fixed and can be described consistently across presentation of information, evaluation of
policies, and valuation of ecological systems and services. In other cases, the physical or

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temporal scale may be variables of interest, so researchers may want to measure whether these
features impact respondents' evaluation of the policy. This could be accomplished by
manipulating the physical or temporal scales of a proposed policy (either between- or within-
subjects) to determine whether and how these features impact support for the policy.

Transfer issues. The most effective way to use surveys for valuation applicable to a
particular ecosystem is to use a survey tailored specifically to that situation. However, this
requires that time and material resources be devoted each time EPA must complete a value
assessment.. A more efficient approach might be to design studies to test whether the findings
from a survey about one set of environmental conditions can be extrapolated to a different set of
environmental conditions. For example, if a survey measures the ecosystem values affected by
one oil spill, would it be possible to multiply these losses by three to anticipate the comparable
losses caused by three comparable oil spills to three comparable ecosystems? Even if such
transformations must be done using more complex transformations, it may be possible to conduct
parametric research to ascertain how such predictions can be made.

Implementing survey research at EPA. Whatever the value measure being sought, the
design and conduct of surveys is best done when informed by the literatures on survey methods.
Therefore, it is important that EPA surveys be implemented at least partly by individuals who are
well-versed and up-to-date in these literatures. This is probably best accomplished by teams of
researchers composed partly of EPA employees who specialize in surveys and outside
consultants who are experts in survey methods. EPA may therefore want to assess its current
capacity to conduct or oversee contractor design and implementation of high-quality surveys.

OMB clearance is required for all EPA surveys, and achieving this clearance requires that
a survey meet high standards of quality. In order to maximize the likelihood of approval, it is
important that a proposed survey meet a set of criteria: a) representative sampling of the
population of interest with minimal non-coverage error; b) a very high response rate or a plan to
assess the presence of non-response bias; c) a measuring instrument that has been developed
according to optimal design and pretesting practices; and d) a measurement approach for which a
body of empirical evidence documents validity.

Probability sampling is relatively easy to do for general population samples, but more
challenging for smaller, more specific subpopulations which require specialized sampling
procedures currently under development (e.g., Blair & Blair, 2006; Rocco, 2003). If EPA is

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interested in conducting surveys of such specialized subpopulations, it may be of value to
commission a group of sampling statisticians to develop a series of guidelines that can be
consulted and followed when conducting sampling for such studies.

The recent literature on response rates has focused on: a) exploring the impact of
response rates on data accuracy, and b) exploring the effectiveness of various data collection
techniques for enhancing response rates. Although lower response rates are generally not
associated with substantially decreased accuracy, it may be useful for EPA to reanalyze a set of
its own past surveys simulating lower response rates and observing the impact on the survey
results. If systematic bias is detected, it may be possible to build correction algorithms to adjust
the results of future surveys to correct for such bias.

It might seem obvious that when EPA conducts surveys, all possible steps should be
taken to increase response rates. According to federal convention, that cannot include offering
financial incentives to respondents, but EPA can implement other techniques to enhance
response rates, including lengthening the field period during which data are collected, and more
attempts to contact potential respondents. However, to justify resources to implement such
techniques, it is important to have empirical evidence documenting the effectiveness of these
techniques for EPA surveys. It is also important to be sure that efforts to increase the response
rate of a survey do not inadvertently decrease the representativeness of the sample. For example,
telling respondents that a survey is about the environment may increase response rates among
people interested in the environment and may decrease response rates by a smaller margin
among less-interested people, thus increasing nonresponse bias. So EPA may want to conduct
studies assessing whether efforts to increase response rates unintentionally decrease sample
representativeness.

Another approach to facilitating OMB approval may be to gather evidence documenting
the effectiveness of particular measurement techniques. For example, there is considerable
controversy surrounding the use of contingent valuation (CV) methods in surveys. Yet NOAA's
Blue Ribbon Panel concluded that CV is a viable method of valuation. It may be of value for
EPA to identify the optimal elements and implementation of a CV survey and to assess the
validity of CV measurement in surveys by comparisons with other monetary measures (e.g.,
from revealed preference studies) or with measures based on judgments of preference,
importance, or acceptability. This same sort of developmental work can be conducted with other

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valuation techniques such as conjoint analysis, about which there is little consensus (e.g., Dennis,
1998; Stevens, et al., 2000; Wainright, 2003). This may help to reassure OMB evaluators of the
merit of value measurements produced by the various methods when they are implemented well.
EPA could also consider conducting research comparing the validity of value assessments by
these and other techniques to identify the technique(s) that yield the most valid data.

Finally, new OMB guidelines on surveys suggest that when a survey is expected to obtain
a relatively low response rate, investigators should plan to implement techniques to assess
sample representativeness. Rather than outlining what such procedures would look like, OMB
has left it to investigators to propose and justify such techniques. EPA could therefore
commission work to design procedures for this purpose and conduct studies to validate the
effectiveness of the procedures.

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ENDNOTES

1	Laws include: the Clean Air Act, Clean Water Act, Comprehensive Environmental
Response, Compensation, and Liability Act, Federal Insecticide, Fungicide and Rodenticide
Act, Toxic Substances Control Act, and Resource Conservation and Recovery Act

2	Although C-VPESS was initiated by the SAB, Senior EPA managers supported the concept
of this SAB project and participated in the initial background workshop that launched the
work of the C-VPESS.

3	The SAB Staff Office published a Federal Register Notice on March 7, 2003 (68 FR 11082-
11084) announcing the project and called for the public to nominate experts in the following
areas: decision science; ecology; economics; engineering; psychology; and social sciences
with emphasis in ecosystem protection. The SAB Staff Office published a memorandum on
August 11, 2003 documenting the steps involved in forming the new committee and
finalizing its membership.

4	The committee developed the conclusions in this report after multiple public meetings and
workshops: a) an Initial Background Workshop on October 27, 2003 to learn the range of
EPA's needs for science-based information on valuing the protection of ecological systems
and services from managers of EPA Headquarters and Regional Offices; b) a Workshop on
Different Approaches and Methods for Valuing the Protection of Ecological Systems and
Services, held on April 13-14, 2004; c) an advisory meeting focused on support documents
for national rulemakings held on June 14-15, 2004; d) an advisory meeting focused on
regional science needs, in EPA's Region 9 (San Francisco) Office on Sept. 13, 14, and 15,
2004; e) advisory meetings held on January 26-26, 2005 and April 12-13, 2005 to review
EPA's draft Ecological Benefits Assessment Strategic Plan; and f) a Workshop on Science
for Valuation of EPA's Ecological Protection Decisions and Programs, held on December 13-
14, 2005 to discuss the integrated and expanded approach described in this paper. The also
committee discussed text drafted for this report at public meetings on October 25 2005; May
9, 2006; October 5-6, 2006, and May 1-2, 2007 and public teleconferences on(insert
additional dates).

5	Likewise, this definition would not include goods or services like recreation that are
produced by combining ecological inputs or outputs with conventional inputs (such as labor,
capital, or time). In addition, Boyd and Banzhaf advocate defining changes in ecosystem
services in terms of standardized units or quantities, which requires that they be measurable
in practice. Such an approach is consistent with the concept of "green accounting," which
extends the principles embodied in measuring marketed products to the measurement and
consideration of the production, or changes in the stock, of ecological or other environmental
"products" (reference NRC report by Nordhaus).

6There is controversy over the meaning of intrinsic value (Korsgaard, C. (1996). Two
Distinctions in Goodness. Creating the Kingdom of Ends. C. Korsgaard. Cambridge,
Cambridge University Press. 1996: 249-74. Many people take intrinsic value to mean that
the value of something is inherent in that thing. Some philosophers have argued that value or
goodness is a simple non-natural property of things (see Moore 1903 for the classical

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statement of this position), and others have argued that value or goodness is not a simple
property of things but one that supervenes on the natural properties to which we appeal to
explain a thing's goodness (this view is defended by, among others, contemporary moral
realists; see McDowell, J. (1985). Values and Secondary Qualities. Morality and Obiectivitv.
T. Honderich, Routledge and Kegan Paul: 110-29., Sturgeon, N. (1985). Moral Explanations.
Morality. Reason, and Truth. D. C. a. D. Zimmerman, Rowman and Allenheld: 49-78; Sayre-
McCord, G. (1988). The Many Moral Realisms. Essays on Moral Realism. G. Sayre-
McCord. Ithaca, Cornell University Press: 1-26;, Brink, D. O. (1989). Moral Realism and the
Foundation of Ethics. Cambridge, Cambridge University Press.

7	One of these elements is an evaluation of willingness to pay for or willingness to accept a
proposed regulatory action and the main alternatives identified and the related costs. The
circular explicitly defines benefits using the economic/utilitarian concept of willingness to
pay (or willingness to accept). The circular contains general guidance on how to provide
monetized, quantitative, and qualitative information to characterize contributions to human
welfare as fully as possible.

8	Under GPRA, the Office of Management and Budget (OMB) requires EPA to periodically
identify its strategic goals and describe both the social costs and budget costs associated with
them. EPA's Strategic Plan for 2003-2008 described the current social costs and willingness-
to-pay or willingness-to-accept analyses of EPA's programs and policies under each strategic
goal area for the year 2002 (U.S. Environmental Protection Agency 2003). This analysis
repeatedly points out that EPA lacks data and methods to quantify willingness to pay or
willingness to accept associated with the goals in its strategic plan. In addition, GPRA
established requirements for assessing the effectiveness of federal programs, including the
outcomes of programs intended to protect ecological resources. EPA must report annually on
its progress in meeting program objectives linked to strategic plan goals and must engage
periodically in an in-depth review [through the Program Assessment Rating Tool (PART)] of
selected programs to identify their net contributions to human welfare and to evaluate their
effectiveness in delivering meaningful, ambitious program outcomes. Characterizing
ecological contributions to human welfare associated with EPA programs is a necessary part
of the program assessment process.

9	These interviews were conducted by one committee member, Dr. James Boyd, in
conjunction with the Designated Federal Officer, Dr. Angela Nugent, over the period
September 22, 2004 through November 23, 2005. In seven sets of interviews, Dr. Boyd
spoke with staff from the Office of Policy, Economics and Innovation, Office of Water,
Office of Air and Radiation, and the Office of Solid Waste and Emergency Response.

10	NCEE is typically brought in by the program offices to both help design and review RIAs.
NCEE can be thought to provide a centralized "screening" function for rules and analysis
before they go to OMB. NCEE is actively involved in discussions with OMB as rules and
supporting analysis are developed and advanced.

11	In addition, the Circular states (p.27) "If monetization is impossible, explain why and
present all available quantitative information" and "If you are not able to quantify the effects,
you should present any relevant quantitative information along with a description of the

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unquantified effects, such as ecological gains, improvements in quality of life, and aesthetic
beauty" (add page number).

12	The Committee reviewed and critically evaluated the CAFO Environmental and Economic
Benefits Analysis at its June 15, 2004 meeting. As stated in the Background Document for
SAB Committee on Valuing the Protection of Ecological Systems and Services for its
Session on June 15, 2004, the purpose of this exercise was "to provide a vehicle to help the
Committee identify approaches, methods, and data for characterizing the full suite of
ecological 'values' affected by key types of Agency actions and appropriate assumptions
regarding those approaches, methods, and data for these types of decisions." The Committee
based its review on EPA's final benefits report (EPA 2002) and a briefing provided by the
EPA Office of Water staff. During the June meeting, members of the Committee divided
into two workgroups. The workgroups each worked independently and reported their
findings to the combined Committee. The leaders of the two working groups then prepared a
consolidated summary of comments from the two workgroups.

13	In December 2000, EPA proposed a new CAFO rule under the federal Clean Water Act to
replace 25-year-old technology requirements and permit regulations (66FR 2959). EPA
published its final rule in December 2003 (68 FR 7176). The new CAFO regulations, which
cover over 15,000 large CAFO operations, reduce manure and wastewater pollutants from
feedlots and land applications of manure and remove exemptions for stormwater-only
discharges.

14	Prior to publishing the draft CAFO rule in December 2000, EPA spent two years preparing
an initial assessment of the costs and benefits of the major options. After releasing the draft
rule, EPA spent another year collecting data, taking public comments, and preparing
assessments of new options. EPA published its final assessment in 2003. An intra-agency
team at EPA, including economists and environmental scientists in the Office of Water,

Office of Air and Radiation, Office of Policy Economics and Innovation, and Office of
Research and Development, worked on the benefit assessment. EPA also worked with the
U.S. Department of Agriculture in developing the assessment. Dr. Christopher Miller of
EPA's Office of Water estimated that EPA spent approximately $1 million in overall contract
support to develop the benefit assessment. EPA spent approximately $250,000-$300,000 on
water quality modeling as part of the assessment.

15	The potential "use" benefits included in-stream uses (commercial fisheries, navigation,
recreation, subsistence, and human health risk), near-stream uses (non-contact recreation,
such as camping, and nonconsumptive, such as wildlife viewing), off-stream consumptive
uses (drinking water, agricultural/irrigation uses, and industrial/commercial uses), aesthetic
value (for people residing, working, or traveling near water), and the option value of future
services. The potential "non-use" values included ecological values (reduced
mortality/morbidity of certain species, improved reproductive success, increased diversity,
and improved habitat/sustainability), bequest values, and existence values.

16	These benefits were recreational use and non-use of affected waterways, protection of
drinking water wells, protection of animal water supplies, avoidance of public water
treatment, improved shellfish harvest, improved recreational fishing in estuaries, and reduced
fish kills.

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17	These include reduced eutrophication of estuaries; reduced pathogen contamination of
drinking water supplies; reduced human and ecological risks from hormones, antibiotics,
metals, and salts; improved soil properties from reduced over-application of manure; and
"other benefits".

18	EPA apparently conducted no new economic valuation studies (although a limited amount
of new ecological research was conducted) and did not consider the possible benefits of
developing new information where important benefits could not be valued in monetary terms
based on existing data.

19	For example, while the report notes the potential effects of discharging hormones and other
pharmaceuticals commonly used in CAFOs into drinking water sources and aquatic
ecosystems, the nature and possible ecological significance of these effects is not adequately
developed or presented. Similarly, the report does not adequately address the well-known
consequences of discharging Trihalomethane precursors into drinking-water sources.

20	In the case of this CAFO rule, 97% of the monetized benefits arise from recreation
(boating, swimming and fishing) and from private well owners' willingness to pay for water
quality, estimated using contingent valuation or travel cost methods.

21	EPA used estimates based on a variety of public surveys in its benefit transfer efforts,
including: a national survey (1983) that determined individuals' willingness to pay for
changes in surface water quality relating to water-based recreational activities (Section 4 of
the CAFO Report); a series of surveys (1992, 1995, 1997) of willingness to pay for
reduced/avoided nitrate (or unspecified) contamination of drinking water supplies (Section
7); and several studies (1988, 1995) of recreational fishers' values (travel cost, random utility
model) for improved/protected fishing success related to nitrate pollution levels in a North
Carolina estuary (Section 9).

22	Although EPA later prepared more detailed conceptual models of the CAFO rule's impact
on various ecological systems and services, EPA did not prepare these models until after the
Agency finished its analysis.

23	Contamination of estuaries, for example, might negatively affect fisheries in the estuary (a
primary effect) but might have an even greater impact on offshore fisheries that have their
nurseries in the estuary (a secondary effect).

24	The goal of EPA's analysis was a national level assessment of the effects of the CAFO
rule. This involved the effects of approximately 15,000 individual facilities, each
contributing pollutants across local watersheds into local and regional aquatic ecosystems. A
few intensive case studies were mentioned in the report and used to calibrate the national
scale models (e.g., NWPCAM, GLEAMS), but there was no indication that these more
intensive data sets were strategically selected or used systematically for formal sensitivity
tests or validations of the national-scale model results.

25	This could include either a robust public involvement process following Administrative
Procedures Act requirements (e.g., FR publication), or some other public involvement

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process [see EPA's public involvement policy, U.S. Environmental Protection Agency Office
of Policy, E. a. I. (2003). Public Involvement Policy of the U.S. Environmental Protection
Agency. EPA 233-B-03-002.; the SAB report on science and stakeholder involvement U.S.
Environmental Protection Agency Science Advisory Board (2001). Improved Science-Based
Environmental Stakeholder Processes: An EPA Science Advisory Board Commentary. EPA-
SAB-EC-COM-OO1-006.

26	Typically production functions in economics have been studied in the context of
businesses that purchase inputs and sell outputs in markets.

27	Models may be valuable in many of the steps of assessing ecological value including:
estimating stress loading; estimating the exposure pattern of stress - especially spatial and
temporal implication; identifying ecological element(s) receiving exposure; estimating
exposure - response function of ecological elements; estimating the reduction or prevention
of increased stress from agency action; estimating the response of service production or
function to change in stress; valuating the ecological service associated with that change in
production; linking to economic or socio-political for further valuation in dollars or other
metrics

28	In theory, one can value a final product either directly (output valuation) or indirectly as
the sum of the derived value of the inputs (input valuation), but not both, since separately
valuing both intermediate and final products leads to double counting. In some cases, it may
be easier or more appropriate to value the intermediate service, while in other cases the
change in the final product can be directly valued.

29	Note that these essential ecosystem characteristics are very similar to the seven ecological
indicators in EPA's report on assessing ecological systems: landscape condition, biotic
condition, chemical and physical characteristics, ecological processes, hydrology and
geomorphology and natural disturbance regimes (Young and Sanzone 2002).

30 One issue relates to the assumption regarding functional form used in the original analysis.
To illustrate the role of functional form, these estimates are interpreted as measuring the
marginal willingness to pay for small improvements in air quality in Chicago. In these
contexts, examples of economic benefits transfer would involve adapting the estimated
marginal willingness to pay (MWTP) for air quality in Chicago so it could be used for
another city, such as Cleveland, New York City, or Los Angeles. If the hedonic price
function used in the Chicago study were linear, the estimated coefficient for the measure of
air quality would be the estimate of the MWTP, which would be constrained to be constant
by the use of the linear price function. In this case, the only adjustment that would be needed
would be for the year of the Chicago study in relationship to the year the analysis sought to
measure the MWTP. Alternatively, if the Chicago analysis used a nonlinear price function,
the MWTP would not be constant and could not be determined solely by the estimated
coefficient of the hedonic price function; rather, the MWTP estimate is itself a function of
variables in the hedonic price function that might be assumed to influence how changes in air
quality affect housing prices. In this case, the adjustment that is needed to conduct an
economic benefit transfer might involve using different values for air quality and other

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determinants of the MWTP that would be associated with the city being studied. It is
important to note that this adjustment is based on an estimate of the MWTP at a particular
point. In particular, it does not assume that the original study estimated a complete
relationship between all air quality levels and housing prices (i.e., a complete marginal
willingness-to-pay function). Estimation of the complete relationship requires added
information. (For a discussion of the distinction between an estimate of MWTP that varies
with other factors versus an estimate of the MWTP function, see Palmquist, 2005).

31	These examples are taken from Ready et al. (2004).

32	An unpublished analysis and peer review of the methods has been developed as part of the
rulemaking process.

33	This complex question reverses the logic used in the conventional analytical framework
used to define an economic benefit measure. An economic benefit measure specifies
something an individual would give up to obtain more of something else. In most
applications, income is the commodity exchanged for a change in some other factor that is
constraining an individual's choices. To assure the definition is complete requires specifying
values for all the other factors that constrain the individual's choices as well as the level of
income and the level of the factor to be relaxed prior to any change.

34	For a more detailed discussion of the sources and possible typologies of uncertainty, see
Krupnick, Morgenstern, et al. (2006).

35	The discussion of value in the National Research Council report (2001) and SAB review of
the EPA's Draft Report on the Environment (US EPA SAB 2005) and related literature (e.g.,
Failing and Gregory, 2003) tends to focus more on qualitative rather than quantitative
expressions. However, issues of scale and aggregation are important. Both the NRC report
(2001) and the SAB review of the EPA's Draft Report on the Environment (U.S. EPA SAB
2005) emphasize the importance of using regional and local indicators. Oover-aggregating
information can obscure critical ecological threats or problems. In general, allowing
sensitivity analysis on disaggregated data is desirable if the data are aggregated at a regional
or higher level. So while some authors recommend simple summary indicators (e.g., Schiller
et al., 2001; Failing and Gregory, 2003), others emphasize disaggregating indicators (U.S.
EPA SAB 2003)

36	This analysis evaluated the benefits and costs of amendments to the Clean Air Act passed
by Congress in 1990. Its effort to evaluate the ecological benefits of these amendments
raises many of the same issues that arise in evaluating the benefits of national rules. In the
prospective analyses the sequence of increasingly stringent rules called for under the 1990
Clean Air Act Amendments are compared with a situation where the rules were held constant
at their 1990 levels (e.g. with the regulatory regime prior to the amendments).

37	A syndrome has been identified that involves: increased biomass of phytoplankton, shifts
in phytoplankton to bloom-forming species that may be toxic, in marine environments,
increases of gelatinous zooplankton, increases in biomass of benthic and epiphytic algae,
changes in macrophytic species composition, decreases in water transparency, oxygen
depletion, increased incidence of fish kills, and loss of desirable fish species (Carpenter et al.,
1998). There are a number of important features of this syndrome. It is easily recognized, it

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is reversible, and there are some features that show up early and hence provide indicators of
ecosystem disruption and early opportunities for mitigation. Clean water and recreational
opportunities have been extensively treated in valuation projects. The impacts on the
biological nature of a system may not be readily appreciated or valued by the public, but it
certainly provides an indicator that the things they do value are in trouble. The power of
public involvement in understanding, valuing and responding to eutrophication is shown by
the classic example from Lake Washington (Smith, 1998). The understanding part took
considerable efforts in educating the public by those few scientists who understood what was
happening.

38	A number of the gasses emitted from CAFOs have adverse air quality impacts that are
interrelated with the water quality impacts.

39	The pollutants that result from CAFOs have environmental effects that are local, regional
and global. For example, in terms of emitted gases, methane and N20 are major greenhouse
gasses of global concern; ammonia and nitrogen oxides have important regional impacts on
air quality and nitrogen deposition; and odor and suspended particulate matter have important
local or on-site impacts (NRC,2003)

40	In the case of air, nitrous oxide has a lifetime of 100 years in the planetary boundary layer,
whereas hydrogen sulfide has a lifetime of only about a day. These spatial and temporal
dimensions of dispersion and lifetime of effects also apply to many of the water pollutants
although the spatial dimensions do not extend to the global.

41	CAFOs are not uniformly distributed in the country or even within a state. For various
reasons they often are clustered. Each of these concentration areas has unique climatic, soil
and topographic features that influence waste dispersion. Further, manure type, in addition to
soil characteristics, has a differential impact on soil microbial populations and hence on
decomposition rates (Larkin 2006).

42	The animal feed used at CAFOs no longer comes from local surroundings but may be
produced in areas remote from the sink facilities, including foreign sources. The production
of these grain feeds results in non-point pollution in the production regions. Further, fish
meal is an important feed supplement for pigs and chickens with the fish generally being
harvested from coastal and marine ecosystems, often from places far distant from the United
States, with consequences for local food chains.

43

Table summarizing Major Chicago Wilderness Reports and Chronology of Valuation Effort

Deci si on/document

Date

Source/URL

Biodiversity Recovery Plan

1999 (Award from
APA in 2001 for best
plan)

http://www.chicasowilderness.o
rs/DubDrod/b rc/index. cfm
Executive summary available at
httD://www.chicasowilderness.o
rs/pubprod/brppdf/CWBRP ch
aoterl.Ddf

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Chicago Wilderness Green
Infrastructure Vision

Final report, March
2004

http://www.nipc.org/environme
nt/sustainable/biodiversity/gree
ninfrastructure/Green%20Infrast
ructure%20 Vi sion%20Final%2
0Report.pdf

Green Infrastructure Mapping



http://www.greenmapping.org/

A Strategic Plan for
the Chicago Wilderness
Consortium

17 March 2005

http://yosemite.epa.gov/SAB/sa
bcvpess.nsf/06347c93513bl813
85256dbf00541478/72clb26a9d
2087568525713f005832el !Ope
nDocument

Chicago Wilderness Regional
Monitoring Workshop
final report by Geoffrey Levin

February 2005

http://yosemite.epa.gov/SAB/sa
bcvpess.nsf/06347c93513bl813
85256dbf00541478/8c33ee9115
d706e68525713f005784e6!0pe
nDocument

Center for Neighborhood
Technology (CNT) - green
infrastructure valuation calculator

Copyright 2004-2007

http : // sreenvalues. cnt. ore/cal cul
ator

44	Consumer surplus measures the excess of the sum of the marginal values over the
expenditures that must be made to obtain the good at a fixed price. Thus, consumer surplus
sums up the differences between the maximum a consumer would be willing to pay for a
good minus the amount actually paid (price) for each unit consumed. Similarly, producer
surplus measures the excess of receipts for the good over the sum of the marginal costs to
provide each unit. Producer surplus is then a comparable concept. It aggregates the difference
between what producers are willing to sell a product for (supply) and what they actually
receive (price) for each unit they provide. Adding together changes in consumer surplus and
producer surplus generates the change in total economic benefit.

45	The last component of these costs, the cost of time on site per visit, is difficult to include
because it is reasonable to assume it is jointly determined with decisions about the location to
visit and the number of trips to take in a season . It is also related to measures of the amount
of the site's services that are consumed. Most studies acknowledge these costs as an issue but
don't include them in the analysis as a result of these difficulties. As a rule the time on site
per trip is assumed to be held constant.

46 The U.S. federal government is one of the largest producers of survey data, which form the
basis of many government policy-making decisions (see Table 1 for examples of federal
funded surveys).

Table 1: Examples of Federal Survevs

Continuously Funded Survevs

Asencv Sponsor

Years

Survey of Income and Program Participation

Census Bureau

1984-
present

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Consumer Expenditure Surveys

Census Bureau

1968-
present

Survey of Consumer Attitudes and Behavior

National Science Foundation

1953-
present

Health and Nutrition Examination Surveys

National Center for Health Statistics

1959-
present

National Health Interview Survey

National Science Foundation

1970-
present

American National Election Studies

National Science Foundation

1948-
present

Panel Study of Income Dynamics

National Science Foundation

1968-
present

General Social Survey

National Science Foundation

1972-
present

National Longitudinal Survey

Bureau of Labor Statistics

1964-
present

Behavioral Risk Factor Surveillance System

Centers for Disease Control and
Prevention

1984-
present

Monitoring the Future

National Institute of Drug Abuse

1975-
present

Continuing Survey of Food Intake by
Individuals

Department of Agriculture

1985-
present

National Aviation Operations Monitoring
System

National Aeronautics and Space
Admin.

2002-
present

National Survey of Drinking and Driving

National Highway Traffic Safety
Admin.

1991-
present

National Survey of Family Growth

National Center for Health Statistics

1973-
present

National Survey of Fishing, Hunting, and
Wildlife-Associated Recreation

Census Bureau

1991-
present

National Survey of Child and Adolescent
Well-Being

Department of Health and Human
Services

1997-
present

Survey of Earned Doctorates

National Science Foundation

1958-
present

National Survey on Drug Use and Health

Department of Health and Human
Services

1971-
present

Youth Risk Behavior Surveillance System

Department of Health and Human
Services

1990-
present

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National Crime Victimization Survey

Bureau of Justice Statistics

1973-
present

Schools and Staffing Survey

National Center for Educational
Statistics

1987-
present

Educational Longitudinal Survey

National Center for Educational
Statistics

2002-
present

Current Employment Statistics Survey

Bureau of Labor Statistics

1939-
present







Other Maior Federallv-Funded Survevs

Agencv Sponsor

National Survey of Distracted and Drowsy
Driving

National Highway Traffic Safety Administration

National Survey of Veterans

Department of Veteran Affairs

National Survey of Children's Health

Health Resources and Services Administration's
Maternal and Child Health Bureau

National Survey of Recent College Graduates

National Science Foundation

National Survey of Speeding and Other
Unsafe Driving Actions

Department of Transportation

47	The use of surveys has also been growing in the private sector and the academic world
(Presser, 1984; Saris, et al., 2003), which likely reflects that (1) surveys are now capable of
generating much more interesting data, via implementation of multifactorial experimental
designs and complex measurement procedures, (2) cross-national comparisons are of
increasing interest, and (3) social scientists want to collect data on more heterogeneous and
representative samples. There is also substantial evidence that the quality of optimally-
collected survey data are generally quite high. For example, in the Monthly Survey of
Consumer Attitudes and Behavior, a representative national sample of American adults has
be